Conservación
Knowledge shortfalls and the effect of wildfireson biodiversity conservation in Guanajuato, Mexico
Tania Escalante a, *, Michelle Farfán b, Oscar Campos a, Leticia M. Ochoa-Ochoa c, Karen Flores-Quintal a, Diego R. García-Vélez a, Ana L. Medina-Bárcenas a, Fernando Saenz a
a Universidad Nacional Autónoma de México, Facultad de Ciencias, Grupo de Biogeografía de la Conservación, Circuito Exterior s/n, Ciudad Universitaria, Coyoacán, 04510 Mexico City, Mexico
b Universidad de Guanajuato, Campus Guanajuato, División de Ingenierías, Departamento de Ingeniería Geomática e Hidráulica, Av. Juárez No. 77, Zona Centro, 36000 Guanajuato, Guanajuato, Mexico
c Universidad Nacional Autónoma de México, Facultad de Ciencias, Departamento de Biología Evolutiva, Museo de Zoología “Alfonso L. Herrera”, Circuito Exterior s/n, Ciudad Universitaria, Coyoacán, 04510 Mexico City, Mexico
*Corresponding author: tescalante@ciencias.unam.mx (T. Escalante)
Received: 01 August 2023; accepted: 29 February 2024
Abstract
Knowledge of shortfalls could modify the geographic distribution patterns and limit the actions to conserve the biodiversity, even in the taxa best known. In addition, forest fires also could modify those patterns, but the potential effects of both factors have not been tested. Our aim was to analyze the effect of the Linnean and Wallacean shortfalls in the first evaluation of wildfire impacts on 22 amphibian and 13 mammal species distributed in Guanajuato, Mexico. We evaluated those shortfalls using the non-parametric estimator Chao2 and the Qs estimator and through maps of species richness patterns. To evaluate the effects of wildfires, we produced a fire recurrence map and quantified the burned area within species distributions and in 24 Protected Natural Areas (PNA) in the state. The Linnean shortfall showed some species missing to record in Guanajuato for both taxa, while the Wallacean shortfall showed poor quality of knowledge. Fire recurrence was high within 5 PNA. The richness patterns affected by fires covered nearly 17% of the surface of Guanajuato. Improving the knowledge of biogeographical patterns could provide better tools to stakeholders to decrease the negative impact of fires within PNA.
Keywords: Fire; Patterns; Priorities; Richness; Species distribution models
© 2024 Universidad Nacional Autónoma de México, Instituto de Biología. Este es un artículo Open Access bajo la licencia CC BY-NC-ND
(http://creativecommons.org/licenses/by-nc-nd/4.0/).
Déficits de conocimiento y el efecto de los incendios forestales en la conservación de la biodiversidad en Guanajuato, México
Resumen
Los déficits en el conocimiento podrían modificar los patrones de distribución geográfica y limitar las acciones para conservar la biodiversidad, incluso en taxones bien conocidos. Además, los incendios forestales también pueden modificar esos patrones, pero los efectos potenciales de ambos no han sido probados. Nuestro objetivo fue analizar el efecto de los déficits Linneano y Wallaceano en la primera evaluación de los impactos de los incendios forestales en 22 especies de anfibios y 13 de mamíferos en Guanajuato, México. Evaluamos esos déficits utilizando los estimadores Chao2 y Qs y con mapas de riqueza de especies. Para evaluar los efectos de incendios forestales, elaboramos un mapa de recurrencia de incendios y cuantificamos el área quemada dentro de las distribuciones de las especies y en 24 áreas naturales protegidas (ANP). El déficit Linneano mostró que faltan algunas especies por registrar para ambos taxones, mientras que el déficit Wallaceano mostró una mala calidad de conocimiento. La recurrencia de incendios fue alta dentro de 5 ANP. Los patrones de riqueza afectados por los incendios cubrieron cerca de 17% de la superficie de Guanajuato. Mejorar el conocimiento de los patrones biogeográficos brindará mejores herramientas para disminuir el impacto de los incendios dentro de las ANP.
Palabras clave: Fuego; Patrones; Prioridades; Riqueza; Modelos de distribución de especies
Introduction
Terrestrial vertebrates are among the best known taxonomic groups, and it is assumed that their distributional areas and their biogeographic patterns are equally well known. However, there are shortfalls that could mask the distributional patterns and therefore, bias the actions to conserve those patterns. Linnean and Wallacean shortfalls affect our knowledge and lead to inaccurate representations of the species richness patterns in taxonomic groups that are presumed to be well known. The Linnean shortfall refers to the discrepancy between formally described species and the number of species that actually exist, while the Wallacean shortfall is the lack of knowledge about the geographical distribution of the species (Hortal et al., 2015; Lomolino, 2004). Both the Linnean and the Wallacean shortfalls can be difficult to evaluate, and they are rarely quantified in the literature prior to a conservation prioritization analysis. For example, to our knowledge, they have never been taken into account when analyzing the effects of wildfires on biodiversity.
Throughout the history of the Earth, fire has been a natural process that has driven the configuration of ecosystems and the maintenance of biodiversity around the world (He et al., 2019; Kelly et al., 2020). Indeed, there are many terrestrial ecosystems that are prone to fire and whose composition and structure are controlled by fire, leading to their classification as fire-adapted ecosystems (He et al., 2019; Schlisky et al., 2007). However, the forest fire regime has been altered by human dynamics associated with fire management and land use change at local and global scales (Chuvieco et al., 2008; Farfán et al., 2018; Martínez-Torres et al., 2015). Several authors agree that there is an increase in the occurrence of wildfires globally (Kelly et al., 2020). Places that did not burn naturally are now burning; examples include the tropical forests of Southeast Asia (Chisholm et al., 2016) and South America (Barlow et al., 2020) to the tundra of the Arctic Circle (Hu et al., 2015). Given the magnitude at which fires are occurring, it has even been proposed that the current era should be coined the Pyrocene, the “age of fire” (Pyne, 2021). This has led to the current situation in which frequency and intensity of forest fires pose a threat to biodiversity conservation worldwide and to human societies; this is due to the damage they cause but also by contributing to global warming.
In Mexico, the effects of wildfires on the fauna are poorly documented. Salazar et al. (2019) proposed a map (scale 1:50,000) of the severity of the fires in the state of Guanajuato for 2017, 2018 and 2019, by calculating the area of burned forest within each of 3 degrees of damage severity: low moderate, high moderate and high. They estimated the total burned area in Guanajuato at 8,460 ha in 2017; 19,589 ha in 2018; and 52,713 ha in 2019 (Salazar et al., 2019). Recently, Farfán et al. (2021) produced a map predicting the occurrence of fires in Guanajuato based on climatic variables under ENSO conditions using a spatial model. They observed that wildfires do not occur in random locations; rather, they are more likely to occur when fragmented forest is immersed in an agricultural matrix, as is frequently the case in the southern part of the state (Farfán et al., 2021).
Figure 1. Location of the state of Guanajuato, Mexico, showing the biogeographic provinces.
Amphibians have been recognized as the most threatened terrestrial vertebrate class (Luedtke et al., 2023), and mammals could have significant declination in their populations due to fire, like in Australia (Geary et al., 2023). However, responses of animals to fire could be diverse because they are strongly related to their life-history traits (González et al., 2021). The responses of the amphibian species are variable and incompletely understood (Pilliod et al., 2003); while even among mammalian species, the effect of fire could be not consistent (González et al., 2021). Therefore, the effect of wildfires on distributional patterns of these taxa could be different.
In this study, our aim is to analyze the Linnean and Wallacean shortfalls in the context of the first evaluation of wildfire impacts on biodiversity. We use species of amphibians and mammals of Guanajuato, Mexico, as a study model and use the results to propose priority areas for conservation.
Material and methods
The state of Guanajuato is located in central Mexico, at 19°55’- 21°51’ N, 99°40’ – 102°06’ W. Most of the state is in the Chihuahuan Desert province, with a few areas in Sierra Madre Oriental and Transmexican Volcanic Belt provinces (Morrone et al., 2017; Fig. 1). Guajanuato has 24 Protected Natural Areas (PNA) mainly located in the southwest and center of the state. The largest of these areas is a biosphere reserve (Sierra Gorda), and the other PNA have a range of different levels of protection and activities allowed (SMAOT, 2022).
Although Guanajuato is not among the most biodiverse states in Mexico, it harbors a wide variety of ecosystems, from Pinus and Quercus forests to xerophytic scrubs (Conabio, 2012). Unfortunately, the extent of the agricultural and livestock areas, the high population density, the expansion of urban areas, and industrial activities have contributed to the destruction and disappearance of the original vegetation (Conabio, 2012).
Guanajuato harbors a total of 27 reported species of amphibians and 93 species of mammals, 8 and 25 of which, respectively, have been described as conservation priorities (DOF, 2010; Leyte-Manrique et al., 2022; Sánchez et al., 2016). Of all the species of amphibians and mammals inhabiting Guanajuato, we chose 22 species of amphibians and 13 species of mammals (Table 1) based on the following criteria: 1) valid nomenclature and at least one point record located in Guanajuato in the consulted databases (Flores-Villela & Ochoa-Ochoa, 2020; Escalante et al., 2018; GBIF.org, 2020a-ah); 2) geographic distribution mostly contained within Mexico; 3) at least 25 point records throughout the whole distribution in order to generate reliable species distribution models (SDM); and 4) considered conservation priorities.
In order to evaluate the Linnean and Wallacean shortfalls for those species, we searched the aforementioned databases for all valid point records in Guanajuato. These were initially overlapped to a grid of 0.25º latitude × 0.25º longitude, in QGIS v.3.16.16 (QGIS Development Team, 2020), which will be referred to hereafter as the “state scale”. We built a presence-absence matrix using the generated SDM.
To evaluate the Linnean shortfall based on point records, we quantified the observed richness (Sobs) as the recorded number of species of amphibians and mammals in each quadrat of 0.25º based on locality records. Then, we used the non-parametric estimator Chao2 in R (Kindt & Coe, 2005), to estimate the asymptotic richness of the incidence dataset (Gotelli & Colwell, 2011). The estimated richness Sest in Chao2 was obtained by the formula (Colwell & Coddington, 1994): Sest = Sobs + L2/(2M), where L = number of species that occur in only 1 quadrat, and M = number of species that occur in exactly 2 quadrats. Thus, this calculation provided a measure of how well the richness of each taxonomic group is known for those areas. We also performed the same analysis based on the species distribution models to explore how well the Linnean shortfall was corrected.
Table 1
List of species of amphibians and mammals in Guanajuato, Mexico, and data used in the analyses. Total records = number of point records after the nomenclatural and geographic validation (state scale). Records in Guanajuato = number of point records of each species into the geopolitical boundaries of the state of Guanajuato. Filtered records = subset of point records after the filter of 10 km applied to the total records. Records for modeling = subset of the filtered records used for model training. pROC = ROC partial of the best model.
| Species | Total records | Records in Guanajuato | Filtered records | Records for modeling | pROC |
| Amphibians | |||||
| Ambystoma velasci | 259 | 45 | 247 | 23 | 1.81 |
| Anaxyrus compactilis | 561 | 31 | 488 | 41 | 1.72 |
| Anaxyrus punctatus | 3,798 | 29 | 3,405 | 226 | 1.76 |
| Aquiloeurycea cephalica | 258 | 2 | 227 | 18 | 1.74 |
| Craugastor augusti | 599 | 16 | 508 | 54 | 1.58 |
| Dryophytes arenicolor | 2,514 | 138 | 2,269 | 161 | 1.81 |
| Dryophytes eximius | 820 | 19 | 761 | 72 | 1.75 |
| Eleutherodactylus guttilatus | 136 | 20 | 109 | 12 | 1.79 |
| Eleutherodactylus nitidus | 673 | 5 | 556 | 48 | 1.68 |
| Eleutherodactylus verrucipes | 237 | 2 | 167 | 13 | 1.52 |
| Hypopachus variolosus | 1,323 | 10 | 1,122 | 116 | 1.74 |
| Incilius nebulifer | 5,585 | 13 | 5,200 | 256 | 1.83 |
| Incilius occidentalis | 1,620 | 81 | 1,342 | 114 | 1.71 |
| Isthmura belli | 245 | 1 | 232 | 18 | 1.64 |
| Lithobates berlandieri | 3,747 | 51 | 3,401 | 247 | 1.56 |
| Lithobates megapoda | 127 | 9 | 101 | 12 | 1.65 |
| Lithobates montezumae | 696 | 94 | 598 | 51 | 1.75 |
| Lithobates neovolcanicus | 349 | 51 | 298 | 32 | 1.68 |
| Lithobates spectabilis | 544 | 3 | 416 | 40 | 1.64 |
| Rheohyla miotympanum | 383 | 1 | 329 | 27 | 1.60 |
| Smilisca baudinii | 4,381 | 2 | 3,590 | 277 | 1.68 |
| Spea multiplicata | 2,251 | 38 | 2,008 | 196 | 1.79 |
| Total | 31,090 | 661 | 27,374 | 2,054 | |
| Mammals | |||||
| Choeronycteris mexicana | 578 | 5 | 384 | 72 | 1.71 |
| Corynorhinus mexicanus | 204 | 4 | 149 | 28 | 1.79 |
| Dipodomys ornatus | 85 | 2 | 61 | 12 | 1.78 |
| Table 1. Continued | |||||
| Species | Total records | Records in Guanajuato | Filtered records | Records for modeling | pROC |
| Leptonycteris nivalis | 263 | 1 | 179 | 34 | 1.66 |
| Leptonycteris yerbabuenae | 575 | 9 | 370 | 70 | 1.75 |
| Lepus callotis | 199 | 6 | 150 | 28 | 1.75 |
| Peromyscus melanotis | 601 | 3 | 227 | 42 | 1.84 |
| Peromyscus difficilis | 897 | 10 | 453 | 85 | 1.69 |
| Peromyscus melanophrys | 594 | 25 | 372 | 70 | 1.59 |
| Rhogeessa alleni | 55 | 5 | 46 | 8 | 1.60 |
| Sciurus oculatus | 123 | 34 | 73 | 14 | 1.45 |
| Sigmodon leucotis | 103 | 9 | 72 | 14 | 1.50 |
| Sorex saussurei | 168 | 1 | 93 | 18 | 1.8 |
| Total | 4,445 | 114 | 2,629 | 495 |
To quantify the Wallacean shortfall, we used the Qs estimator (Murguía-Romero & Villaseñor, 2000), which is a measure of the quality of the records. QS can take values between ‘0’ and ‘1’ and is defined as (Murguía-Romero & Villaseñor, 2000): QS = F/[Sobs m/(1 – Es) – max (Sobs, m)], where F = the sum of frequencies of all classes multiplied by all classes (that is, the sum of all ‘1’ in the matrix); Es = measure of the proportion of the known richness related to the estimated richness; and m = the total number of quadrats. Murguía-Romero and Villaseñor (2000) characterized QS values above 80% as “very good”, values between 50% to 80% as “good” and less than 50% as “poor” data quality.
To compare the possible effect of the Wallacean shortfall in the biogeographic patterns, we performed a species distribution model (SDM) for each species in order to obtain a map of richness patterns for amphibians and mammals. Following the BAM diagram of Soberón and Peterson (2005), where the M corresponds to the region that is reachable by the species from established distributional areas in ecological time (Soberón & Peterson, 2005). The M for each species was obtained using the concept of extent of occurrence, defined as “the area contained within the shortest continuous imaginary boundary that can be drawn to encompass all the known, inferred or projected sites of present occurrence of a taxon, excluding cases of vagrancy” (IUCN, 2001). Therefore, we defined M as the area within a minimum convex hull polygon for each species constructed in QGIS v.3.16.16 (QGIS Development Team, 2020).
The M of each species was used to crop the 19 environmental layers of WorldClim 2 (Fick & Hijmans, 2017) and 3 topographic variables (slope, elevation and aspect; USGS, 2021) at ~ 1 km2 of resolution. The data points were filtered in Wallace software (Kass et al., 2018) to a distance of 10 km between points to reduce spatial biases, and retain useful information (Aiello-Lammens et al., 2015; Pearson et al., 2007). To avoid collinearity among the 22 variables for each species, we obtained the VIF (Mandeville, 2008; Montgomery & Peck, 1992), applying the packages usdm (Naimi et al., 2014) and rgdal (Bivand et al., 2015) in RStudio (RStudio Team, 2020).
The models were performed in the maximum entropy package kuenm (Cobos et al., 2019) in RStudio (RStudio Team, 2020). The occurrence dataset for each species was divided as follows: 75% of the points were used for training and 25% for testing; and a set of independent occurrences of 25% for a last evaluation; those datasets were built with the kuenm_occsplit function in kuenm. For the next step, we used the function kuenm_cal, using the feature classes: linear, quadratic and hinge; and the regularization multipliers 0.5, 1, 1.5, 2, 2.5, 3, 3.5 and 4. All models were evaluated with kuenm_ceval, calculating the ROC partial with E = 10 (Peterson et al., 2008), and Akaike criterion for small samples (AICc; Warren & Seifert, 2011). The final best model for each species was obtained on a clog-log scale using the pROC value in NicheToolbox (Osorio-Olvera et al., 2020).
To produce binary maps of geographic distribution area of each species, the final best model was reclassified using the “10th percentile training presence” threshold, and cropped to the political boundaries of Guanajuato.
Wildfires. We used hotspot data from the MODIS sensor at a resolution of 1 km2 for the years 2000 to 2021, downloaded from NASA Earth Data Cloud (2020). Each hotspot was overlapped to a net of ~ 1 km2 covering the state of Guanajuato in QGIS (QGIS Development Team, 2020), and we counted the number of hotspots in each square of the net. This area will be referred to as the “fine scale”.
In order to compare the number of fires in each square of the net with the scale of the models, we transformed this number between ‘0’ and ‘1’ through a min-max normalization (Farrús et al., 2007; Jain et al., 2005). This procedure was useful to evaluate the effect of the wildfires on the SDM of each species. All cartographic products were projected to UTM zone 14 north coordinates, which corresponds to the state of Guanajuato. To quantify the impact of the fires on species richness patterns, we rasterized the map of the number of fires, from which we produced a new map of kernel density with a radius of 3,000 m, using the software DINAMICA EGO (Ferreira et al., 2019). The map of kernel density was multiplied by the richness map for each taxonomic group. Finally, we quantified the burned area for the maps and for the Protected Natural Areas (SMAOT, 2022).
Results
Shortfalls. We obtained 31,090 records of amphibians considering the whole distribution for the 22 amphibian species, and 661 records within Guanajuato. For the 13 species of mammals, the number of records was 74 in Guanajuato and 2,629 in the whole distribution. The species with the most records was the coastal plain toad Incilius nebulifer (5,585 records) for the amphibians and the southern rock mouse Peromyscus difficilis (453 records; see Table 1) for the mammals. However, within Guanajuato, there were only 13 and 9 records for these species, respectively. The species with the highest number of point records within Guanajuato were the canyon tree frog Dryophytes arenicolor with 138 records, and the Peters’s squirrel Sciurus oculatus, with 34 records.
The richest 0.25º quadrat for amphibians had 11 species and the quadrat with the most mammals had 5 species, while the lowest number of species per quadrat was 1 for both taxa, although there were a few marginal quadrats without data for amphibians and some complete quadrats without data for mammals (~ 12). The maps of quadrats with the observed richness (Sobs in Chao2) for both groups are shown in the Supplementary material: figures 1S, 2S. The main results of the Chao2 estimator are shown in Table 2. For both groups, the number of observed species (Sobs) was lower than the expected number (Sest) for Guanajuato. In the case of the Chao2 estimated with species distribution models, the species richness was the same as the expected species.
Regarding the Wallacean shortfall, the Qs estimator had a value of 23% for the amphibians and 28% for mammals at the state scale. These values were categorized as “poor” quality data in both taxa, but Qs was worse in amphibians.
Table 2
Results of the recorded and estimated richness for the complete distributional data of amphibians and mammals of Guanajuato, Mexico based both on species richness and on species distribution models (SDM).
| Taxonomic group | Recorded richness (Sobs) | Estimated richness (Sest) | Standard deviation (SD) | Estimated richness with SDM (Sest) |
| Amphibians | 22 | 24.61 | 3.42 | 24 |
| Mammals | 13 | 14.94 | 3.64 | 14 |
Figure 2. Richness pattern for 22 modeled species of amphibians and the Protected Natural Areas in Guanajuato, Mexico (black polygons): 1. Palenque, 2. Peña Alta, 3. Sierra de Pénjamo, 4. Sierra de los Agustinos, 5. Las Fuentes, 6. Sierra de Lobos, 7. Las Musas, 8. Lago Cráter La Joya, 9. Cerro de los Amoles, 10. Parque Metropolitano, 11. Cerro de Arandas, 12. Cuenca Alta del Río Temascatio, 13. Cerro del Cubilete, 14. Cerros El Culiacán y La Gavia, 15. Cuenca de La Esperanza, 16. Mega Parque de la Ciudad de Dolores Hidalgo, 17. Presa La Purísima y su Zona de Influencia, 18. Presa de Neutla y su Zona de Influencia, 19. Laguna de Yuriria y su Zona de Influencia, 20. Pinal de Zamorano, 21. Región Volcánica Siete Luminarias, 22. Sierra Gorda de Guanajuato, 23. Cuenca de la Soledad, 24. Presa de Silva y Áreas Aledañas.
Some results of the modeling process are shown in Table 1. The distribution models of amphibians predicted that the species Craugastor augusti, Dryophytes arenicolor, Dryophytes eximius, Incilius occidentalis, Lithobates montezumae, Lithobates neovolcanicus, and Spea multiplicata are distributed in more than 90% of the surface of Guanajuato. Meanwhile, the species Aquiloeurycea cephalica (31%), Smilisca baudini (25%), Lithobates berlandieri (19%), and Incilius nebulifer (3%) had the lowest proportion of distribution in Guanajuato. Regarding the mammals, Choeronycteris mexicana, Corynorhynus mexicanus, Leptonycteris nivalis, Lepus callotis, Peromyscus difficilis, Peromyscus melanophrys, and Peromyscus melanotis were the most widely distributed in Guanajuato (more than the 90% of the state is predicted as part of their distribution), while Sorex saussurei was the only species with a proportion less than 30%.
For the richness patterns, the pixels with highest number of species modeled to be present were similar between amphibians and mammals at the fine scale (Figs. 2, 3), showing a diagonal strip of high richness from the northwest to the southeast, which coincided with 14 PNA: Las Fuentes, Sierra de Lobos, Parque Metropolitano, Cuenca Alta del Río Temascatío, Cerro del Cubilete, Cuenca de La Esperanza, Presa de Neutla, Cuenca de la Soledad, Sierra de los Agustinos, Lago Cráter La Joya, Cerro de los Amoles, Cerros El Culiacán y La Gavia, Laguna de Yuriria and Región Volcánica Siete Luminarias. There were also other sites of high diversity, for example within the PNA of Sierra de Pénjamo and Sierra Gorda.
Wildfires. The maximum number of wildfires per quadrant at the fine scale was 6. To further explore the risk of wildfire recurrence, we built a risk map using the recurrence of fires in each square, with 3 classes: 1) low risk, for pixels with one fire during the analyzed period; 2) medium risk, for pixels with 2 or 3 fires; and 3) high risk, for pixels with 4, 5 or 6 fires (Fig. 4). A large proportion of quadrats with high recurrence of fires occurred outside PNA (for example at northern Guanajuato), but there were also some high risk zones within PNA, like Palenque, Peña Alta, Sierra de Pénjamo, Las Musas and Región Volcánica Siete Luminarias.
Figure 3. Richness pattern for 13 modeled species of mammals and the Protected Natural Areas in Guanajuato, Mexico (black polygons): 1. Palenque, 2. Peña Alta, 3. Sierra de Pénjamo, 4. Sierra de los Agustinos, 5. Las Fuentes, 6. Sierra de Lobos, 7. Las Musas, 8. Lago Cráter La Joya, 9. Cerro de los Amoles, 10. Parque Metropolitano, 11. Cerro de Arandas, 12. Cuenca Alta del Río Temascatio, 13. Cerro del Cubilete, 14. Cerros El Culiacán y La Gavia, 15. Cuenca de La Esperanza, 16. Mega Parque de la Ciudad de Dolores Hidalgo, 17. Presa La Purísima y su Zona de Influencia, 18. Presa de Neutla y su Zona de Influencia, 19. Laguna de Yuriria y su Zona de Influencia, 20. Pinal de Zamorano, 21. Región Volcánica Siete Luminarias, 22. Sierra Gorda de Guanajuato, 23. Cuenca de la Soledad, 24. Presa de Silva y Áreas Aledañas.
Figure 4. Map of fire risk in Guanajuato between the years 2000 and 2021, based on the recurrence of fires in a 1 km square, with 3 classes: (1) low risk, for squares with one fire during the analyzed period; (2) medium risk, for squares with 2 or 3 fires; and (3) high risk, for squares with 4, 5 or 6 fires. Red polygons represent the Protected Natural Areas.
Table 3
Potential distribution area predicted by the modeling for 22 amphibian and 13 mammal species and quantification of habitat lost due to wildfires relative to the total area of Guanajuato, Mexico (30,702 km2).
| Species | Surface of potential distribution area occupied in Guanajuato (km2) | Percentage of potential distribution area occupied (%) | Surface of potential distribution area affected by wildfires (km2) | Percentage of potential distribution area affected by wildfires (%) |
| Amphibians | ||||
| Ambystoma velasci | 22,351 | 73 | 3,620 | 16 |
| Anaxyrus compactilis | 26,061 | 85 | 4,948 | 19 |
| Anaxyrus punctatus | 18,241 | 59 | 2,440 | 13.3 |
| Aquiloeurycea cephalica | 9,653 | 31 | 1,246 | 12.9 |
| Craugastor augusti | 30,702 | 100 | 5,264 | 17.1 |
| Dryophytes arenicolor | 30,696 | 99.9 | 5,262 | 17.1 |
| Dryophytes eximius | 30,267 | 98.5 | 5,249 | 17.3 |
| Eleutherodactylus guttilatus | 21,636 | 70 | 2,088 | 14.2 |
| Eleutherodactylus nitidus | 22,888 | 75 | 4,663 | 20.3 |
| Eleutherodactylus verrucipes | 20,519 | 67 | 2,756 | 13.4 |
| Hypopachus variolosus | 19,076 | 62 | 3,951 | 20.7 |
| Incilius nebulifer | 873 | 3 | 55 | 6.3 |
| Incilius occidentalis | 30,546 | 99 | 5,257 | 17.2 |
| Isthmura bellii | 26,647 | 87 | 4,925 | 18.4 |
| Lithobates berlandieri | 5,772 | 19 | 614 | 10.6 |
| Lithobates megapoda | 20,866 | 68 | 4,422 | 21.1 |
| Lithobates montezumae | 29,957 | 97.5 | 5,242 | 17.4 |
| Lithobates neovolcanicus | 28,670 | 93 | 5,137 | 17.9 |
| Lithobates spectabilis | 24,821 | 81 | 3,893 | 15.6 |
| Rheohyla miotympanum | 18,056 | 59 | 3,516 | 19.4 |
| Smilisca baudinii | 7,631 | 25 | 1,579 | 20.6 |
| Spea multiplicata | 30,661 | 99.8 | 5,264 | 17.1 |
| Mammals | ||||
| Choeronycteris mexicana | 30,608 | 99.69 | 5,242 | 17.12 |
| Corynorhynus mexicanus | 30,608 | 99.69 | 5,242 | 17.12 |
| Dipodomys ornatus | 18,553 | 60.43 | 3,429 | 18.48 |
| Leptonycteris nivalis | 30,608 | 99.69 | 5,242 | 17.12 |
| Leptonycteris yerbabuenae | 24,955 | 81.28 | 4,474 | 17.93 |
| Lepus callotis | 28,469 | 92.73 | 5,025 | 17.65 |
| Peromyscus difficilis | 30,142 | 98.18 | 5,225 | 17.34 |
| Peromyscus melanophrys | 29,606 | 96.43 | 5,227 | 17.65 |
| Peromyscus melanotis | 29,231 | 95.21 | 5,183 | 17.73 |
| Rhogessa alleni | 17,604 | 57.34 | 3,353 | 19.05 |
| Sciurus oculatus | 14,564 | 47.44 | 1,923 | 13.20 |
| Sigmodon leucotis | 18,215 | 59.33 | 3,139 | 17.23 |
| Sorex saussurei | 4,151 | 13.52 | 756 | 18.22 |
Figure 5. Richness map of amphibians intersected with the kernel density of recurrence of the wildfires in Guanajuato, Mexico. Black polygons represent the Protected Natural Areas.
1. Palenque, 2. Peña Alta, 3. Sierra de Pénjamo, 4. Sierra de los Agustinos, 5. Las Fuentes, 6. Sierra de Lobos, 7. Las Musas, 8. Lago Cráter La Joya, 9. Cerro de los Amoles, 10. Parque Metropolitano, 11. Cerro de Arandas, 12. Cuenca Alta del Río Temascatio, 13. Cerro del Cubilete, 14. Cerros El Culiacán y La Gavia, 15. Cuenca de La Esperanza, 16. Mega Parque de la Ciudad de Dolores Hidalgo, 17. Presa La Purísima y su Zona de Influencia, 18. Presa de Neutla y su Zona de Influencia, 19. Laguna de Yuriria y su Zona de Influencia, 20. Pinal de Zamorano, 21. Región Volcánica Siete Luminarias, 22. Sierra Gorda de Guanajuato, 23. Cuenca de la Soledad, 24. Presa de Silva y Áreas Aledañas.
Respect to the temporal distribution of the recurrence of forest fires, for the period of time analyzed, the years 2017, 2019, and 2021 had the highest number of fires (72, 110 and 150, respectively). The map of kernel density is shown in Supplementary material: Figure 3S. The map was transformed to a binary map and overlapped with the patterns of richness of amphibians and mammals, to produce the maps in figures 5, 6.
The evaluation of the effects of the wildfires on the model of each species is shown in Table 3. The mean percentage of burned potential distribution area was 16.49 and 17.37 for amphibians and mammals, respectively. Some species’ distribution areas were more strongly affected by fires, such as Eleutherodactylus nitidus, Hypopachus variolosus, Lithobates megapode, and Smilisca baudinii, all of which were amphibians for which more than 20% of their distribution area had been burned. For mammals, the most affected species was Rhogessa alleni (19.05%), followed by Sorex saussurei (18.22%).
The area of Guanajuato affected by fires measured nearly 5,200 km2 (17%; Figs. 5, 6), with high diversity zones for amphibians located in the northwest, near Sierra de Lobos (PNA 6) and Peña Alta (PNA 2). Other important affected areas for amphibians coincided with high recurrence of wildfires in the southern of Guanajuato in Región Volcánica Siete Luminarias (PNA 21), Sierra de los Agustinos (PNA 4) and Cerro de los Amoles (PNA 9). For mammals, some of the most strongly affected areas coincided with those of the amphibians (e.g., within Cerro de los Amoles; PNA 9), but there were also areas that were unique to mammals (Fig. 6). For example, there were areas of high mammal richness with wildfires in central Guanajuato, which did not coincide with any PNA, as well as southeastern areas in Sierra de los Agustinos (PNA 4). Fortunately, areas with high richness for both amphibians and mammals were not affected by fires, like Sierra de Lobos (PNA 6), Cuenca de la Esperanza (PNA 15), and Cuenca de la Soledad (PNA 23).
Discussion
Linnean shortfalls in the state of Guanajuato could have medium effects because the estimator predicted at least 2 additional species for each taxon relative to total currently known. This suggests that the current species inventories are not yet complete. This finding does not dismiss possible shortfalls at more detailed scales, because the number of records within the state of Guanajuato is very low, with an average of 30 per each species of amphibian and only 6 of each species of mammal. It would therefore be informative to carry out more specific analyses within the quadrats where 0 or 1 species were recorded. Increased collection effort in the field could improve the problems of undersampling, since the number of total data points in Guanajuato is very low for some species (v. gr. Aquiloeurycea cephalica with 2 records, and Leptonycteris nivalis with 1 record; Table 1). It is interesting to highlight that when performing the analyses with SDM, the Linnean shortfall is apparently corrected (Table 2). However, these results should be taken with caution, because it is possible that there are commission errors in models or that those areas actually correspond to sister species (Rodrigues et al., 2019; Acevedo et al., 2014).
Figure 6. Richness map of mammals intersected with the kernel density of recurrence of the wildfires in Guanajuato, Mexico. Black polygons represent the Protected Natural Areas.
1. Palenque, 2. Peña Alta, 3. Sierra de Pénjamo, 4. Sierra de los Agustinos, 5. Las Fuentes , 6. Sierra de Lobos, 7. Las Musas, 8. Lago Cráter La Joya, 9. Cerro de los Amoles, 10. Parque Metropolitano, 11. Cerro de Arandas, 12. Cuenca Alta del Río Temascatio, 13. Cerro del Cubilete, 14. Cerros El Culiacán y La Gavia, 15. Cuenca de La Esperanza, 16. Mega Parque de la Ciudad de Dolores Hidalgo, 17. Presa La Purísima y su Zona de Influencia, 18. Presa de Neutla y su Zona de Influencia, 19. Laguna de Yuriria y su Zona de Influencia, 20. Pinal de Zamorano, 21. Región Volcánica Siete Luminarias, 22. Sierra Gorda de Guanajuato, 23. Cuenca de la Soledad, 24. Presa de Silva y Áreas Aledañas.
On the other hand, the Wallacean shortfall was highly relevant for both amphibians and mammals, showing poor quality. Murguía-Romero and Villaseñor (2000) suggested that the quality of the records is related to the geographical resolution of the biogeographical analysis. In future analysis within the state of Guanajuato and using smaller quadrats (for example, close in size to the pixels of our models), the Wallacean shortfall could strongly affect the observed data, reaching very poor data quality. Thus, it seems to be the more important shortfall for these vertebrate species.
For our 25 species, probably the effect of the Linnean shortfall has a less dramatic effect than the Wallacean shortfall. There are multiple potential explanations for these shortfalls, including low intensity and spatial variation of sampling, which can directly affect biodiversity estimators like species richness (Oliveira et al., 2016). Continuing the study of these shortfall will be important for the correct implementation of conservation strategies, for example with other methods including correlations using the sampling effort (Oliveira et al., 2016), rate of descriptions and number of taxonomists (Joppa et al., 2011), many different algorithms of species distribution modeling and maps of ignorance (Oliveira et al., 2016; Rocchini et al., 2011; Tessarolo et al., 2021), among others.
In spite of the extant Wallacean shortfalls, the richness patterns for both taxa were partially recognized; specially, the recorded most richness quadrats at 0.25º also showed the modeled pattern for amphibians. For mammals, the highest richness area near to Cuenca de la Esperanza was identified also for a quadrat of 0.25º, coinciding with the models. Amphibians and mammals shared many (though not all) areas of high richness, generally following a northwest-southeast diagonal across the state. The partial similarity in those richness patterns could be useful in Systematic Conservation Planning, because both taxa could represent each other as good surrogates (Escalante et al., 2020).
In general, amphibians and mammals are over-represented taxa in databases, although amphibians are less represented than mammals (Troudet et al., 2017). Few articles have quantified the Linnean and Wallacean shortfalls prior to biogeographic analysis. In particular, Oliveira et al.(2016) suggested that terrestrial vertebrates have similar biases compared with some taxa of arthropods, contradicting the statement that terrestrial vertebrates are better suited for biogeographic and conservation studies. In some places, such as the state of Guanajuato, amphibians and mammals could have similar Linnean shortfalls, but differ in the severity of their Wallacean shortfalls, which could modify the biogeographic patterns identified.
The relevance of including shortfall analysis in biogeographical studies, mainly in those related to species conservation, lies in the fact that the Linnean and Wallacean shortfalls strongly influence the possible results, since the data on the identity and distribution of the species are crucial to identify patterns in biodiversity, as well as the processes that modify those patterns (Hortal et al., 2015). In particular, Wallacean shortfalls can also alter estimates of threatened conservation status, since range size is regularly used in conservation (Hortal et al., 2015). Species with small ranges have higher priority in many international and national standards (DOF, 2010; IUCN, 2012). Therefore, Wallacean shortfalls could lead to some taxa and areas being disproportionately prioritized over others because their distribution areas have been erroneously underestimated (Riddle et al., 2011). In addition, other shortfalls that potentially can affect the biogeographic patterns should be investigated, like Darwinian shortfall (Diniz-Filho et al., 2013, 2023), and even distinct categories of Linnean shortfalls (Vergara-Asenjo et al., 2023).
As we expected, wildfires affected all species, but in different ways. For amphibians, the species Smilisca baudinii, in addition to having a small distribution area in Guanajuato compared to the rest of the species, is one of the most affected by wildfires occurrences. These observations may suggest that Smilisca baudinii should be considered a priority species for conservation in the state of Guanajuato. Furthermore, we also highlight Lithobates megapoda, which is listed in the 2019 update of the NOM-059-SEMARNAT-2010 (DOF, 2019) as a species under special protection and described as sensitive to habitat degradation (Santos-Barrera & Flores-Villela, 2004).
On the other hand, in the northeastern part of the state, the effects seem to be minimal compared to the south and southwest, but it is important to remember that species with specific habitat requirements such as Incilius nebulifer, Lithobates berlandieri, and Smilisca baudinii are distributed in this area. Therefore, attention should be paid to investigating the sources of ignition present at this area in order to prevent future wildfires, since the loss of habitat could result in the disappearance of these species. Finally, in accordance with Clivillé et al.(1997), who describe the effect of fires on amphibians from 3 points of view (habitat, species and individual), our study only focuses on the effect of these events on the habitat. Thus, the effects on the distribution of the selected species can be interpreted as loss of habitat and vegetation cover, trophic resources and humidity due to wildfires, which are determining characteristics for the presence of amphibians and their reproduction (Clivillé et al., 1997).
For the case of mammals, Sorex saussurei (the Saussure’s shrew) is the species with the narrowest geographic distribution in Guanajuato. This shrew is only distributed in Mexico and Guatemala, and even though it is considered as least concern on the Red List of Threatened Species (IUCN, 2017), some populations in Mexico have been categorized as threatened and under special protection (DOF, 2019). Secondly, Peters’s squirrel Sciurus oculatus occupies less than 50% of the surface of Guanajuato and is also categorized as least concern in the Red List of Threatened Species (IUCN, 2016). Sciurus oculatus is found only within Mexico, and is under special protection in national legislation NOM-059-SEMARNAT-2010, update of 2019 (DOF, 2019). Both species face continuing decline in the extent and quality of habitat due to land use change (Conafor, 2020; IUCN, 2016, 2017), which is exacerbated by repeated burning episodes that decrease the area occupied by each species by 13-18%. According to Zamudio (2012), most of the plant communities in the state of Guanajuato have significantly changed in their structure, floristic composition and physiognomy. Consequently, their distribution areas have been gradually reduced. Currently, 63% of the territory has been transformed into agricultural areas, human settlements, and areas devoid of vegetation (Roth et al., 2016).
The recurrence of wildfires, mainly in the southern part of Guanajuato, represents an important threat to biodiversity conservation within PNA, which are surrounded by a complex matrix of rainfed and irrigated agricultural land uses. This result was also found by Farfán et al. (2020, 2021), where both the probabilities of anthropogenic ignition and climate under the ENSO climate conditions lead to high wildfire risk in this region of the state. These PNA urgently need fire management plans that can integrate fire prevention actions at the local level in the context of global warming. On a global scale, wildfires have been responsible for up to 27% of the loss of tree cover between 2001 and 2021 (Tyukavina et al., 2022). In Mexico alone, in 2021, 408.75 km2 of forest were lost due to fire (Tyukavina et al., 2022). The effects of wildfires on biodiversity patterns could be understimated if these shortfalls are underestimating the biodiversity. Therefore, actions at the international level are also urgent in order to prevent damage to unknown biogeographic patterns.
This is the first study for the state of Guanajuato and for Mexico that addresses the effect of wildfire on the potential distribution of 2 important taxonomic groups: amphibians and mammals. The evaluation of the Linnean and Wallacean shortfalls for any taxonomic group is essential before the identification of geographic patterns involved as criteria for conservation planning, even in terrestrial vertebrates which are assumed to be adequately sampled. The Wallacean shortfall could lead to underestimations of the effects of perturbations such as wildfires. This is particularly true of species that are already vulnerable due to anthropogenic factors such as land cover change, illegal trafficking, etc., as well as intrinsic factors like the size of their natural distributional areas, because it is unknown whether undersampling could represent geographically rare species. Improving the biogeographical knowledge of the patterns of amphibians and mammals could provide better tools to stakeholders in order to generate fire management plans to prevent the negative impact of the wildfire within protected areas around the world.
Acknowledgements
We thank Julián A. Velasco and Luis J. Aguirre for their help with the parametrization of Maxent, and Miguel Murgía for assistance with the quantification of the Qs estimator. We also thank to the anonymous reviewers and the Associate Editor for their careful reading of this manuscript.
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Monitoreo poblacional y estado de conservación de la ranita del Pehuenche (Alsodes pehuenche) en el valle Pehuenche, Mendoza, Argentina
Gabriela Diaz a, b, *, Vanesa Pellegrini-Piccini a, Liliana Moreno d, Martín Palma c, e, Vanesa Bentancourt c y Valeria Corbalán f
a Universidad Nacional de Cuyo-Sede Malargüe, Campus Educativo Municipal, Facultad de Ciencias Exactas y Naturales, Rosario Vera Peñaloza y Beltrán, 5613 Malargüe, Mendoza, Argentina
b Universidad Nacional de Cuyo, Instituto de Ingeniería y Ciencias Aplicadas a la Industria-CONICET, Facultad de Ciencias Aplicadas a la Industria, Bernardo de Yrigoyen Núm. 375, 5600 San Rafael, Mendoza, Argentina
c Instituto de Educación Física Núm. 9-016 “Jorge E. Coll” Dirección General de Escuelas-Sede Malargüe, Tecnicatura en Conservación de la Naturaleza, Campus Educativo Municipal, Rosario Vera Peñaloza y Beltrán, 5613 Malargüe, Mendoza, Argentina
d Universidad Nacional de San Luis, Facultad de Química Bioquímica y Farmacia, Ejército de los Andes Núm. 950, 5700 San Luis, Argentina
e Ministerio de Ambiente y Desarrollo Sustentable de la Provincia de Mendoza, Dirección de Recursos Naturales Renovables, Delegación Malargüe, San Martín Norte Núm. 352, 5613 Malargüe, Mendoza, Argentina
f Instituto Argentino de Investigaciones de Zonas Áridas (CCT Mendoza-CONICET), Av. Ruiz Leal s/n, Parque Gral. San Martín, 5500 Mendoza, Argentina
*Autor para correspondencia: gdiaz@infoar.net (G. Diaz)
Recibido: 02 octubre 2023; aceptado: 15 agosto 2024
Resumen
La ranita del Pehuenche, Alsodes pehuenche, es endémica de los Andes centrales de Argentina y Chile, ha sido categorizada en peligro crítico por la UICN y entre sus amenazas se encuentran la ruta internacional que atraviesa los arroyos que habita, la presencia del hongo quitridio, los salmónidos exóticos invasores, el ganado y el cambio climático. El objetivo de este trabajo fue evaluar el estado actual de conservación de A. pehuenche en el valle Pehuenche para conocer tendencias poblacionales, el impacto de las amenazas y futuras acciones de manejo. Se realizaron 14 salidas de campo durante 3 temporadas (2021-2023) y se muestrearon 12 arroyos usando la técnica de encuentro visual nocturno. Se delimitaron y nombraron 7 subpoblaciones: Nacientes, del Límite, Pichintur, Rial Rojas, Nueva, Campanaria y Cajón Largo. Los resultados muestran conteos de adultos (5.82 en 200 m2 y 13.64 por hora) y de larvas (6.24 en 200 m2 y 17.76 por hora). Éstos no variaron significativamente entre temporadas, pero fueron mayores en enero y febrero. Con base en la conectividad y las amenazas, los índices del estado de conservación permiten priorizar las subpoblaciones como unidades de conservación, de las cuales la del Límite requiere esfuerzos más urgentes.
Palabras clave: Especie amenazada; Encuentro visual; Conectividad; Subpoblaciones; Priorización de conservación
© 2024 Universidad Nacional Autónoma de México, Instituto de Biología. Este es un artículo Open Access bajo la licencia CC BY-NC-ND
(http://creativecommons.org/licenses/by-nc-nd/4.0/).
Population monitoring and conservation status of the Pehuenche frog (Alsodes pehuenche) in the valle Pehuenche, Mendoza, Argentina
Abstract
The Pehuenche spiny-chest frog, Alsodes pehuenche, is endemic to the Central Andes of Argentina and Chile. It has been categorized as critically endangered by the IUCN and its threats include the international road that crosses the streams inhabited by the species, the presence of the chytrid fungus, invasive exotic salmonids, livestock, and climate change. The objective of this work was to evaluate the current conservation status of A. pehuenche in the Pehuenche Valley as a basis for understanding population trends, the impact of threats, and future management actions. Fourteen field trips were conducted during 3 seasons (2021-2023) and 12 streams were sampled using the nocturnal visual encounter technique. Seven subpopulations were delimited and named: Nacientes, del Límite, Pichintur, Rial Rojas, Nueva, Campanaria, and Cajón Largo. The results show counts of adults (5.82 in 200 m2 and 13.64 per hour) and larvae (6.24 in 200 m2 and 17.76 per hour). These did not vary significantly between seasons but were higher in January and February. According to connectivity and threats, the conservation status indices allow us to prioritize the subpopulations as conservation units, with del Límite being the one that requires the most urgent efforts.
Keywords: Threatened species; Visual encounter; Connectivity; Subpopulations; Conservation prioritization
Introducción
Los anfibios son el grupo de vertebrados más amenazados de nuestro planeta, varias son las causas responsables de la disminución de sus poblaciones (Grant et al., 2020; Green et al., 2020; Luedtke et al., 2023). Más de 45% de la diversidad de anfibios del mundo se distribuyen en el Neotrópico (Kacoliris et al., 2022). Alrededor de 25% de las especies de Argentina son endémicas (Vaira et al., 2017), 37% de ellas en Argentina y Chile se encuentran en disminución, mientras que del 22% de las especies no se conoce su tendencia poblacional (Kacoliris et al., 2022). Las amenazas más importantes con las que se asocian la disminución poblacional o extinciones locales son los depredadores invasores, enfermedades emergentes y la ganadería (Kacoliris et al., 2022; Velasco et al., 2016).
La ranita del Pehuenche, Alsodes pehuenche, fue descrita por Cei (1976). Luego de estudios citogenéticos realizados entre 1983 y 2003 (Cuevas y Formas, 2003), se renueva el interés sobre la especie en Argentina debido a las obras viales sobre la ruta internacional ARG145 – CH115, cuya pavimentación desvió el curso de 5 afluentes del arroyo Pehuenche con presencia de la especie (Corbalán et al., 2010). La ranita del Pehuenche es una especie endémica y su distribución está restringida a los Andes centrales de Argentina y Chile (Corbalán et al., 2010, 2023; Correa et al., 2013, 2018, 2020). Habita arroyos de montaña en ecosistemas de vegas o mallines entre 2,150 y 2,825 m snm (Corbalán et al., 2023). Los arroyos poseen lechos pedregosos y una fina capa de sedimentos. La especie presenta dimorfismo sexual y como otras especies de Alsodes, tiene larvas de desarrollo prolongado con juveniles y adultos de hábitos acuáticos (Cei, 1976, 1980; Herrera y Velázquez, 2016a; Úbeda, 2021). Las larvas son de gran tamaño y pasan al menos 4 años en los cuerpos de agua permanentes hasta completar el ciclo larval (Corbalán et al., 2014). Se ha reportado la puesta de huevos ocultos bajo rocas o en oquedades en las márgenes de los arroyos (Corbalán et al., 2014; Piñeiro et al., 2020). Estas cavidades son utilizadas también como refugio por los adultos (Cei, 1980; Correa et al., 2013; Herrera y Velázquez, 2016b).
Los datos de conteos disponibles corresponden a muestreos de tramos cortos de arroyos en Argentina y Chile (Corbalán et al., 2010, 2023; Correa et al., 2013). La coexistencia de individuos en una cavidad durante el día sugiere densidades elevadas en una categoría de microhábitat (Correa et al., 2013). Han ocurrido eventos de mortalidad y se han reportado nuevas poblaciones (Corbalán et al., 2023; Correa et al., 2018), lo cual hace necesaria una evaluación actualizada de su estatus de conservación y tendencia poblacional.
Las evaluaciones del estado de conservación de la ranita del Pehuenche la han colocado en la categoría más alta de amenaza: “en peligro” por la Asociación Herpetológica Argentina (Vaira et al., 2012), cuarta en orden de prioridad entre los 58 anfibios evaluados de Chile (Vidal et al., 2024) y “en peligro crítico” por la Unión Internacional para la Conservación de la Naturaleza (IUCN, 2019), según los criterios B1ab basados en su extensión de presencia estimada y su disminución continua estimada, provocada principalmente por la pavimentación de la ruta. El área de ocupación de la especie (AOO sensu UICN) estimada actualmente es de 4.84 km2 y su extensión de presencia (EOO sensu UICN) 497.9 km2 (Corbalán et al., 2023). En cuanto a las amenazas consideradas en la categorización de IUCN (2019), se enumeran el desvío de los cursos de agua por la construcción de la ruta, el impacto del ganado, el cambio climático, la presencia del hongo quitridio (Batrachochytrium dendrobatidis) y la depredación por salmónidos exóticos invasores tales como la trucha arcoíris (Oncorhynchus mykiss)y trucha marrón (Salmo trutta) (Corbalán et al., 2023; Ghirardi et al., 2014; Zarco et al., 2020).
Si bien los avances en el conocimiento de la distribución de A. pehuenche han sido importantes en los últimos años, aún se desconocen aspectos básicos de la ecología, sistemática, reproducción, comportamiento y estrategias ecofisiológicas de esta especie en los humedales de altura. Por tratarse de una especie endémica y amenazada, la historia de vida y demografía son fundamentales para la evaluación del estatus de conservación (Luja et al., 2015). Los programas de seguimiento de poblaciones son necesarios para identificar y detectar disminuciones que amenacen la persistencia de poblaciones y deben llevarse a cabo en un marco de gestión adaptativa que permita realizar monitoreos que maximicen la detección y minimicen el esfuerzo (Pollock, 2006, Yoccoz et al., 2001). A su vez, la definición de subpoblaciones es una herramienta útil que permite definir su estado actual y priorizar acciones de conservación (Velasco, 2018).
El objetivo de este trabajo fue evaluar el estado de conservación actual de A. pehuenche en el valle Pehuenche. Esta información es fundamental para estimar tendencias poblacionales a largo plazo, evaluar el impacto de las amenazas y el éxito de futuras acciones de manejo. En este trabajo se muestran los primeros datos del programa de monitoreo iniciado en 2021 en el valle Pehuenche. A partir del mismo, se definen y delimitan subpoblaciones como unidades de conservación sobre las que se deben priorizar las acciones.
Materiales y métodos
El área de estudio corresponde al valle Pehuenche, en el lado argentino de la zona limítrofe entre Argentina y Chile; forma parte de los Andes centrales, en el departamento de Malargüe, suroeste de la provincia de Mendoza. Las precipitaciones anuales son de 400 a 600 mm (Rivera et al., 2018) y están influidas por los vientos provenientes del Pacífico sur, creando un gradiente de precipitación y humedad de oeste a este con fuertes nevadas en invierno (Garreaud et al., 2009). Durante la primavera-verano, el deshielo alimenta los humedales, denominados localmente vegas o mallines, donde se asientan familias con su ganado en los puestos, denominados reales o riales. El área es considerada un corredor ecológico y cultural trashumante de gran importancia (Llano et al., 2021).
La cuenca del arroyo Chico en Argentina incluye las subcuencas donde se distribuye A. pehuenche: arroyos Pehuenche y Callao (Corbalán et al., 2023). El arroyo Chico es afluente del río Grande en el departamento de Malargüe, provincia de Mendoza. En Chile, el área de distribución de la especie se ubica en la cuenca del río Maule, con 2 subpoblaciones posiblemente aisaldas: laguna del Maule y Lo Aguirre (Correa et al., 2013, 2018).
Durante 3 temporadas se muestrearon 14 arroyos tributarios del arroyo Pehuenche (fig. 1). Doce de los 14 arroyos fueron muestreados exhaustivamente y en los 2 restantes (B0 y A8), solo se constató la presencia de la especie. La especie fue observada en la desembocadura de los arroyos de primer orden en el arroyo Pehuenche.
Siguiendo a Corbalán et al. (2010) y Prado et al. (2019), los arroyos de primer orden que desembocan en el arroyo Pehuenche se distinguen, según su ubicación respecto a este último, en arroyos A que se ubican hacia el sur y son atravesados por la ruta Núm. 145, mientras que los arroyos B se ubican al norte del arroyo Pehuenche (fig. 1). A los arroyos con bifurcaciones y cursos de agua paralelos a menos de 100 m de distancia se los denominó con el mismo nombre. Durante el estudio, se tomaron datos de temperatura del agua con termómetro de mercurio, así como pH, oxígeno disuelto y conductividad con sonda multiparamétrica Lutron WA-2017SD, y de temperatura ambiente mediante la aplicación yr (Jensen et al., 2007). La conductividad fue baja, entre 0 y 181.56 μS/cm (n = 65), el valor medio del pH estuvo cercano a la neutralidad 6.41 ± 1.2, variando entre 3.29 y 8.25 (n = 77) y la concentración de oxígeno disuelto varió entre 1.3 y 15.8 mg/l (n = 16). La temperatura ambiente durante los muestreos osciló entre 2 y 21 ºC (n = 89), y la temperatura del agua entre 2 y 17.6 ºC (n = 75).
Se realizaron un total de 14 salidas de campo con una duración de 2 a 4 días/noches (tabla 1). Las 14 salidas se distribuyeron en 3 temporadas; cada una se inicia con el deshielo (octubre-noviembre-diciembre) y finaliza con la caída de las primeras nevadas (marzo-abril). En enero de 2021 se iniciaron muestreos preliminares diurnos y nocturnos consistentes en encuentros visuales sin manipulación de individuos. Los muestreos nocturnos iniciaron luego de la puesta del sol que coincide con el pico de actividad de los individuos adultos. Se optó por continuar con los muestreos nocturnos debido a que no se obtuvieron diferencias significativas en la cantidad de larvas entre el día y la noche (W = 191, p = 0.859).
Figura 1. Arroyos muestreados en la cuenca del valle Pehuenche, distribuidos en 7 subpoblaciones locales: I. Nacientes, II. del Límite, III. Pichintur, IV. Rial Rojas, V. Nueva, VI. Campanaria, VII. Cajón Largo. Los arroyos de cada subpoblación se indican con el mismo color.
Se contabilizaron adultos, juveniles y larvas usando la técnica de encuentro visual (Crump y Scott, 1994). Los transectos fueron de rumbo variable siguiendo el curso del arroyo y el ancho fue de 1 m a cada lado del arroyo. Cada transecto se realizó con 2 a 5 observadores. Siguiendo las recomendaciones de Pereyra et al. (2021), al menos un observador tenía experiencia previa y las salidas se programaron en días con condiciones meteorológicas similares: días sin lluvias, mayormente despejados y con vientos leves a moderados.
Para los estadios maduros se registró sexo (macho, hembra e indeterminado). Se registró punto GPS y hora de inicio y de finalización del transecto. Se calcularon las abundancias relativas como individuos/área e individuos/tiempo. El área de los arroyos se consideró como el tramo muestreado en metros multiplicado por 2 m de ancho. Los individuos contabilizados estuvieron a no más de 1 m de distancia del curso del arroyo, cuando éste está claramente definido, y solo se observaron más dispersos en las zonas húmedas de vegas. Para el cálculo de individuos/tiempo se usó el tiempo en minutos de la duración del muestreo (Pereyra et al., 2021).
Adicionalmente, se realizaron muestreos en transectos de 800 m de longitud en los márgenes de la calzada: 10 en la temporada estival de 2021 y 5 en las temporadas siguientes. Este registro se realizó ya que en enero de 2018 se observaron individuos deshidratados o muertos, adyacentes a un cordón o bordillo de 20 cm de altura y 15 cm de ancho.
Se definieron subpoblaciones, siguiendo a Velasco (2018), formadas por un grupo de arroyos que contienen individuos de A. pehuenche, con hábitat adecuado (sensu Corbalán et al., 2010; Correa et al., 2013) y cuya distancia entre ellos no fue mayor a 500 m lineales. De este modo, los 14 arroyos iniciales fueron agrupados en 7 subpoblaciones (fig. 1). Los nombres de las subpoblaciones se relacionan con referencias a las características del paisaje donde se encuentran, excepto la subpoblación denominada Nueva. Se calculó el índice de conectividad (IC) entre subpoblaciones de acuerdo con la fórmula de Lin (2009), modificada por Velasco (2018): IC = A1*(1/D1*T1) + A2*(1/D2*T2), donde Ai es el área (= tamaño poblacional) de la población local vecina hacia el lado i, Di es la distancia a la población local vecina hacia el lado i y Ti la presencia de truchas en hábitat intermedios (presencia multiplica por 2, ausencia multiplica por 1). Para estimar el área correspondiente a cada subpoblación, se usaron las coordenadas GPS de los muestreos de campo y en el caso de las 2 subpoblaciones donde solo se constató presencia, se estimó tamaño del área en Google Earth Pro.
Tabla 1
Temporadas y subpoblaciones en las que se realizaron los muestreos de Alsodes pehuenche. Se indican la cantidad de muestreos realizados en la ruta y los nombres de los arroyos (tipos A, B; P: arroyo Pehuenche) muestreados por subpoblación.
| Temporada | Mes | Quincena primera (1) segunda (2) | Ruta | Subpoblaciones | ||||||
| I | II | III | IV | V | VI | VII | ||||
| 1 (2021) | enero | 1 | 2 | B0 | A1, A2 | |||||
| enero | 2 | 6 | A1, A2, A3, A4 | A5 | A7 | |||||
| febrero | 2 | 1 | A1, A2 | B2 | B3 | |||||
| abril | 2 | 1 | A1, A2 | A5 | A8 | |||||
| 2 (2021-2022) | noviembre | 2 | 1 | A1, B1 | B2 | A5 | A7 | |||
| febrero | 1 | A2 | B3 | |||||||
| marzo | 1 | A1 | B2 | |||||||
| marzo | 2 | B1 | B3 | |||||||
| 3 (2022-2023) | noviembre | 2 | 1 | A5 | ||||||
| diciembre* | 1 | 1 | A1, A3, B1 | B2, B2 | A5-B3 | A6 | A7 | |||
| enero | 1 | 1 | A2, P | A5, B3, P | A6, P | A7, P | ||||
| febrero | 2 | 1 | A1, P | A7, P | ||||||
| marzo | 2 | A3, P | B4, P | B6, P |
Las amenazas que se registraron durante los conteos se consideraron de manera cualitativa, en una escala de 0 a 3, de acuerdo con la intensidad de la amenaza (Velasco, 2018). En la tabla 2 se muestran las amenazas consideradas: 1) la presencia de salmónidos que fueron registrados por encuentro visual; 2) la sequía observada de arroyos que podría ser reversible según la época del monitoreo y cárcavas permanentes cuya profundidad va en aumento; 3) la presencia de ganado, principalmente vacuno y caprino, el primero es el que podría causar mayor impacto debido al mayor uso de las vegas; 4) mortalidad de individuos por causas indeterminadas; 5) la ruta, considerada como una barrera para la dispersión, ya sea porque las alcantarillas no fueron diseñadas como pasos de fauna o porque la altura del cordón o bordillo construido sobre la misma es muy alto para el libre tránsito de ranas juveniles y adultas. Y a pesar de que la ruta representa una barrera, la construcción posterior de rampas, con el objetivo de mitigar este impacto, se considera una acción concreta de conservación. Se registró, además, la presencia de depredadores no acuáticos potenciales y de otras especies de anfibios, pero ninguno de estos casos se considera como amenaza, aunque podrían representar interacciones interespecíficas negativas.
El estado de conservación (EC) relativo de cada subpoblación se obtuvo a partir de la sumatoria de los valores asignados al índice de conectividad y las amenazas (Velasco, 2018). Los valores fueron estandarizados (ECE = EC subpoblación – promedio de todos los EC/desviación estándar de todos los EC) y el valor resultante se multiplicó por -1. Así, los valores negativos de ECE quedaron relacionados con situaciones menos favorables. Se establecieron 2 categorías de prioridades de conservación: baja a media (valores positivos de ECE) y alta (valores negativos de ECE). Los valores altos son los que requieren medidas más urgentes de manejo (Velasco, 2018).
Tabla 2
Variables consideradas para evaluar estado de conservación de las subpoblaciones de Alsodes pehuenche en el valle Pehuenche. IC: Índice de conectividad. Valores 0 a 3 son los que se le asignan a cada variable.
| Variables | 0 | 1 | 2 | 3 |
| Área (m2) | 7,500-10,000 | 5,000-7,500 | 2,500-5,000 | 1,000-2,500 |
| IC1 | 7.51-10 | 5.1-7.5 | 2.51-5 | 0-2.5 |
| Amenazas | ||||
| Efecto especies exóticas: salmónidos | sin | presencia registrada | presencia registrada en más de un conteo | presencia registrada y presencia de anfibio/s con signo de depredación |
| Fragmentación por la ruta | Sin ruta | Sin cordón o bordillo | Con cordón o bordillo y con rampas | Con cordón o bordillo y sin rampas |
| Pérdida de hábitat por sequía/ cárcavas | Sin evidencias de sequía ni cárcavas | Algún arroyo parcialmente seco | Algún arroyo parcialmente seco y/o con presencia cárcavas | Algún arroyo completamente seco y/o presencia de individuos muertos por sequía |
| Efecto del ganado | Sin ganado | Con caprinos | Con vacunos | Con vacunos y cabras |
| Indeterminada | Sin registro de individuos muertos sin causa aparente | Registro de individuos con manchas en la piel | Registro de individuos con manchas y uno muerto sin causa aparente | Registro de más de un individuo muerto sin causa aparente |
1 = Índice de conectividad calculado de acuerdo con Lin (2009), modificado por Velasco (2018).
Se realizaron análisis de los conteos con R versión 4.1.3 y los análisis estadísticos consideraron 0.05 de nivel de significación de alfa. La normalidad de los datos se verificó con la prueba de Shapiro-Wilk. La cantidad de individuos (adultos, juveniles y larvas) por área y por hora no se distribuyó normalmente (W < 0.82934, p < 0.001), por lo que se usaron las pruebas estadísticas no paramétricas de Mann-Whitney-Wilcoxon y de Kruskal-Wallis para 2 grupos y más de 2 grupos, respectivamente. Se evaluó estadísticamente si hubo diferencias entre adultos y juveniles considerados en conjunto versus larvas. También se evaluaron diferencias en relación con variables temporales (temporadas, meses) y espaciales (entre arroyos A y B y entre subpoblaciones). Cuando se obtuvieron diferencias significativas con la prueba de Kruskal-Wallis, se usó el paquete conover.test versión 1.1.5 (2017), basado en Conover-Iman Test para comparaciones múltiples. En todos los casos, los resultados se expresan en mediana, el rango intercuartílico (IQR) como medida de dispersión, los valores mínimos y máximos (min-max) y el tamaño de la muestra (n).
Resultados
Los transectos muestreados alcanzaron un total de 51,292 m lineales recorridos en 181 horas con un esfuerzo de muestreo 12.35 horas/persona. De los 14 arroyos estudiados, todos tuvieron presencia de la especie y el área total fue de 0.04 km². Más de la mitad corresponde a los arroyos denominados A (56.2%).
Cuando se analizaron los conteos respecto de la cantidad de observadores, la correlación de Spearman resultó en valores de rho cercanos a 0, lo que sugiere que no hay correlación lineal (p > 0.05), tanto para individuos postmetamórficos (por área: rho = -0.0427, n = 99; por hora: rho = 0.0975, n = 99), como para larvas (por área: rho = -0.0297, n = 92; por hora rho = -0.0063,
n = 92).
A lo largo del periodo de muestreo se observaron 5.82 adultos y juveniles/área (IQR = 9.27, 0-45 min-max, n = 99) y 13.64 adultos y juveniles/hora (IQR = 30.29, 0-108 min-max, n = 99); 6.24 larvas/área (IQR = 31, 0-175 min-max, n = 92) y 17.76 larvas/hora (IQR = 65, 0-367 min-max, n = 92). La comparación entre estadios muestra menos postmetamórficos que larvas, tanto para los conteos de individuos por área (W = 12.524, p < 0.001) como por hora (W = 26.918, p < 0.001).
Tabla 3
Prioridades de conservación para las subpoblaciones de Alsodes pehuenche en el valle Pehuenche, con base en el tamaño de la población, la conectividad y las amenazas.
| Sub-población | Área (m2) | IC | Amenazas | ECE | PC1 | ||||
| Ruta | Sequía/cárcavas | Ganado | Muertes indet. | Salmónidos | |||||
| I. Nacientes | 4,044 | 8.41 | 0 | 0 | 2 | 0 | 1 | 0.88 | Baja |
| II. del Límite | 5,744 | 8.39 | 2 | 3 | 2 | 2 | 3 | -1.26 | Muy alta |
| III. Pichintur | 5,002 | 9.91 | 0 | 0 | 2 | 3 | 2 | 0.08 | Media |
| IV. Rial Rojas | 6,208 | 6.05 | 1 | 3 | 2 | 1 | 2 | -0.73 | Alta |
| V. Nueva | 2,160 | 9.61 | 1 | 0 | 2 | 3 | 0 | -0.19 | Media |
| VI. Campanaria | 5,860 | 5.26 | 1 | 0 | 1 | 3 | 3 | -0.46 | Media |
| VII. Cajón Largo | 9,550 | 6.64 | 0 | 0 | 1 | 0 | 0 | 1.69 | Muy baja |
1 PC: Prioridad de conservación calculada de acuerdo con Velasco (2018).
Las comparaciones entre las 3 temporadas no mostraron diferencias significativas para ninguno de los grupos estudiados adultos y juveniles/área (H = 1.9935, df = 2, p = 0.369), adultos y juveniles/hora (H = 1.3599, df = 2, p = 0.5066), larvas/área (H = 0.13936, df = 2, p = 0.933) y larvas por hora (H = 0.6051, df = 2, p = 0.7389). Entre meses, la diferencia no fue significativa para larvas (H = 4.5278, df = 5, p = 0.4762 por área, H = 4.0986, df = 5, p = 0.5353 por hora), en cambio para los postmetamórficos solo fue significativa para los conteos por área (H = 13.97, df = 5, p = 0.0158). Los análisis post hoc indican menores cantidades de adultos y juveniles en febrero respecto a diciembre (p = 0.0075) y enero (p = 0.0021), como también en marzo respecto a enero (p = 0.0185, fig. 2).
Figura 2. Variación mensual de la abundancia de adultos y juveniles (individuos/200 m2) de Alsodes pehuenche. Los meses con diferentes códigos de letras indican diferencias significativas (p < 0.05).
Figura 3. Abundancia de Alsodes pehuenche (individuos/200 m2) entre 5 subpoblaciones. A, Adultos y juveniles; B, larvas. Las subpoblaciones con diferentes códigos de letras indican diferencias significativas (p < 0.05).
De los 14 arroyos del valle Pehuenche con presencia de A. pehuenche, los arroyos A tienen un área de ocupación de 49% respecto de los arroyos B con 51%. Y en los 12 arroyos donde se realizaron conteos por área, éstos fueron mayores en los arroyos A que en los B, tanto para adultos y juveniles (H = 31.977, df = 2, p < 0.001), como para larvas (H = 10.42, df = 2, p = 0.0055).
Se reconocieron 7 subpoblaciones en el valle Pehuenche: Nacientes, del Límite, Pichintur, Rial Rojas, Nueva, Campanaria y Cajón Largo (fig. 1). De éstas, se tienen datos de conteo de 5 (todas, excepto Nacientes y Cajón Largo). Cuando se comparan estas 5 subpoblaciones, se encuentran diferencias significativas tanto para adultos y juveniles/área (H = 17.436, df = 4, p = 0.0016), como para larvas/área (por área H = 23.639, df = 4, p < 0.001). Se registraron más adultos y juveniles en las subpoblaciones Rial Rojas y Nueva, y más larvas en Rial Rojas (fig. 3). Sin embargo, estas diferencias no son significativas cuando se analizan los datos de postmetamórficos/hora (H = 8.8053, df = 5, p = 0.1171) y larvas/hora (H = 4.0986, df = 5, p = 0.5353).
El único lugar donde se encontraron solo larvas fue en una charca de la subpoblación del Límite. Éstas coexistían con larvas y adultos de rana de 4 ojos (Pleurodema bufoninum). Además, esta última especie ha sido observada eventualmente coexistiendo con A. pehuenche en las subpoblaciones Rial Rojas, Campanaria y del Límite. Otras especies registradas durante los muestreos, que se identifican como posibles depredadores de A. pehuenche son el zorro gris (Lycalopex gymnocercus), el chiñe o zorrino (Conepatus chinga), la gaviota capucho café (Chroicocephalus maculipennis) y el águila mora (Geranoaetus melanoleucus).
Aunque la detección de huevos no fue objeto de este estudio, dado que no se realizó búsqueda activa en oquedades ni se manipularon hembras, en noviembre 2022 se observó una masa de 20 huevos aproximadamente, flotando en una pequeña poza de la subpoblación Rial Rojas en época de deshielo. Los huevos no eran pigmentados y cada uno midió entre 6 y 7 mm de diámetro.
Los índices de conectividad calculados para cada una de las subpoblaciones se muestran en la tabla 3. El arroyo Pehuenche representa la principal conexión entre subpoblaciones. En este arroyo se encontraron individuos adultos en la desembocadura de arroyos de primer orden pertenecientes a las subpoblaciones del Límite, Rial Rojas, Nueva y Campanaria (tabla 1). Si bien Cajón Largo tiene relativamente bajo índice de conectividad, el área de esta subpoblación es la mayor de todas.
Se registraron 24 individuos muertos durante el periodo de estudio: 1) 6 individuos adultos —2 hembras, 2 machos y 2 adultos indeterminados— en el mes de enero de la temporada estival 2021 y 2 en la temporada 2022-2023 en los muestreos realizados en la ruta internacional ARG145; 2) 3 individuos —1 adulto indeterminado y 2 larvas— encontrados muertos por desecación sobre el curso de un arroyo de Rial Rojas que mostró una disminución abrupta del caudal luego del deshielo entre noviembre y diciembre 2022; 3) 4 adultos —3 machos y 1 indeterminado— encontrados con signos de depredación (patas traseras lastimadas o mutiladas) al costado o dentro del curso del arroyo donde se registra presencia de salmónidos, esto es, en las subpoblaciones del Límite y Campanaria; 4) 6 adultos —2 machos y 4 indeterminados—, 1 juvenil indeterminado y 2 larvas encontrados muertos sin causa evidente, dentro del curso del arroyo o próximo a éste.
Finalmente, los valores de ECE negativos usados como indicadores, permitieron identificar a las subpoblaciones del Límite y Rial Rojas como unidades de conservación prioritarias, de las cuales la del Límite requiere esfuerzos más urgentes.
Discusión
Corbalán et al. (2010) reportaron 355 individuos de A. pehuenche en 2,500 m lineales (equivalente a 14 individuos/200 m2) en un muestreo realizado en 2008 en arroyos de la subpoblación del Límite. Si se compara ese valor con el obtenido en este estudio para los mismos arroyos en el mismo mes en 2021 (8 individuos/200 m2), el primero es mayor, pero se encuentra dentro del intervalo de variación de este estudio. Cabe mencionar que se usaron 2 técnicas diferentes para el conteo (búsqueda activa de individuos en oquedades durante el día y encuentro visual nocturno, con 3 observadores en los 2 casos), por lo que, con la información disponible actualmente, solo la mortalidad de individuos reportada desde 2008 permite estimar una tendencia poblacional decreciente para esta subpoblación.
En 2 arroyos muy cercanos a la subpoblación del Límite en Chile, Correa et al. (2013) reportaron 20 y 24 adultos/h o 15 y 66 adultos/200 m². Estos últimos datos corresponden a muestreos nocturnos realizados en marzo de 2012, con 4 observadores. La cantidad máxima de adultos por área observados por Correa et al. (2013) es mayor al máximo obtenido en este estudio, lo cual puede deberse a un error de muestreo en solo 30 m lineales, o bien a fluctuaciones poblacionales naturales (Kissel et al., 2020). En arroyos de Rial Moreno, a 21 km al NE de la población del valle Pehuenche, Corbalán et al. (2023) reportaron 26 adultos/hora y 4 juveniles/hora en un muestreo nocturno realizado en enero de 2019, los cuales se encuentran dentro de valores obtenidos en el valle Pehuenche en este estudio.
Durante las 3 temporadas consecutivas muestreadas en este trabajo, no se detectaron cambios significativos para ningún estadio de desarrollo de A. pehuenche. Esto puede deberse a que se trata de un periodo corto en términos de monitoreo. Por lo tanto, es necesario mantener los muestreos para conocer tendencias poblacionales a largo plazo. Por otro lado, la detección de individuos por encuentro visual a lo largo de los arroyos, podría compararse con otros estudios enfocados al monitoreo poblacional de la especie para lograr establecer un óptimo en términos de esfuerzo de muestreo. Dado que estas investigaciones dependen del presupuesto disponible y de limitaciones logísticas (Joseph et al., 2006) principalmente ligadas a la elevada demanda de horas/persona en un ambiente de alta montaña, se debe buscar una estrategia que permita satisfacer tanto la demanda de datos pertinentes para la gestión, como la obtención de información ecológica de calidad (Stephens et al., 2015; Yoccoz et al., 2001).
En relación con la primera acción de conservación llevada a cabo en la población del valle Pehuenche, que consistió en la construcción de 800 m de cordón o bordillo en la ruta y 2 cámaras de infiltración para evitar que la sal vertida en la ruta llegue a los arroyos, provocó la deshidratación y muerte de 18 individuos que fueron encontrados en diciembre de 2017, y 50 en enero de 2018. Posiblemente, la construcción a posteriori de rampas entre el cordón o bordillo y la calzada en febrero de 2018, ha evitado que se vuelvan a registrar mortalidades tan abruptas. Si bien nunca pudo constatarse el uso de las rampas por parte de la especie, las mismas habrían mejorado la conexión entre las secciones superior e inferior de los arroyos. Esto demuestra la necesidad de seguimiento de las acciones de conservación que se llevan a cabo. Pero, por otro lado, si bien la ruta pudo haber provocado una disminución poblacional (Corbalán et al., 2010; IUCN, 2019), el número elevado de individuos muertos registrado en este estudio alerta sobre la necesidad de avanzar en el conocimiento de otras amenazas, como la depredación por salmónidos invasores, el cambio climático y la infección por el hongo quitridio (Batrachochytrium dendrobatidis) de los anfibios.
En este trabajo, de todos los estadios evaluados, las larvas fueron las más abundantes, las más estables en el tiempo y presentaron el mismo patrón de variación que los adultos entre arroyos A y B, y subpoblaciones (fig. 3). Al igual que lo reportado por Corbalán et al. (2010, 2023), se encontraron larvas de diferentes tamaños conviviendo en pozas, pero también se observaron en rápidos, remansos y debajo de la vegetación acuática. Estos resultados son coherentes con la estrategia de especies con desarrollo larval plurianual (Úbeda, 2021). Por otro lado, la menor cantidad de juveniles registrada respecto de adultos podría ser un problema de detectabilidad por su tamaño (Petrovana y Schmidt, 2019), podría corresponder a una segregación por microhábitats según el estadio (Gonwouo et al., 2022), o bien, ser evidencia de una dinámica poblacional particular solo en esta fase de su ciclo de vida (Kissel et al., 2020).
Los muestreos preliminares realizados de día y de noche demuestran que hay arroyos permanentes que solo tienen adultos. Los estudios a futuro podrán determinar las causas que posiblemente puedan estar asociadas con aspectos fisicoquímicos del agua, o a la dispersión de individuos. La información actual permite suponer que las rutas de dispersión de A. pehuenche con base en sus hábitos acuáticos, son los mismos arroyos. Los registros de individuos secos en la ruta, principalmente entre diciembre y enero, al igual que la mayor cantidad de adultos y juveniles registrados en estos meses (fig. 2), podrían estar indicando una mayor dispersión en esa temporada del año.
Los arroyos del valle Pehuenche son los mejores conocidos hasta el momento y el área que ocupan representa solo 6% del área de distribución de A. pehuenche (487.9 km²) reportada por Corbalán et al. (2023). De la misma manera, habiendo considerado en nuestro estudio casi todos los arroyos presentes en el valle Pehuenche, representa 0.8% del área de ocupación total para la especie (4.84 km²; Corbalán et al., 2023). El espacio sin ocupación efectiva de la especie, actualmente, es muy grande (Corbalán et al., 2023), aunque en el pasado pudo haber estado presente en un área mayor si se comparan los registros de presencia en Cajón Grande (Cei y Roig, 1965; Corbalán et al., 2023). Si se considera desde la descripción de la especie en 1965, A. pehuenche ha reducido su área de distribución histórica, lo cual puede deberse a diferentes causas. La depredación por salmónidos en la actualidad podría ser la causa de que la especie haya quedado restringida solo a las zonas altas de las cuencas. Por ello, es necesario conocer la distribución de salmónidos e identificar dónde coexisten estos peces y la rana. Donde esto ocurre en el valle Pehuenche, se registraron adultos muertos de A. pehuenche con signos de depredación. Así, los salmónidos representan una barrera para la dispersión de las ranas, reduciendo la conectividad entre subpoblaciones y dentro de ellas entre arroyos A y B (tabla 3). Como se ha demostrado para otras especies de anfibios, la consecuencia es la disminución del intercambio genético o demográfico (e.g., Kacoliris et al., 2022; Velasco, 2018). Por lo tanto, para conservar las subpoblaciones se requiere asegurar su conectividad, tanto entre las secciones inferior y superior de los arroyos A impactados por la ruta, como dentro de las subpoblaciones entre los arroyos A y B, y entre las subpoblaciones conectadas a través del arroyo Pehuenche.
Las observaciones de campo sugieren que los individuos se distribuyen de manera agrupada a lo largo de los cursos de agua y vegas, no solo las larvas en las pozas, sino también los adultos a lo largo de los arroyos. Esto debe ser evaluado en la etapa reproductiva a inicios de la temporada y al final de ésta, cuando el caudal disminuye y algunos cauces se secan total o parcialmente, donde solo quedan pozas pequeñas con agua. Si este efecto es acentuado por la disminución de cobertura permanente de nieve en la cuenca del Río Grande (Aumassanne et al., 2019), es posible que más arroyos, que todavía mantienen agua hasta el final de la temporada, se sequen como ha ocurrido en los arroyos A3 y A4 de la población del Límite, aumentando así la fragmentación del hábitat. Si A4 no se hubiera secado, las 3 subpoblaciones (del Límite, Pichintur y Rial Rojas) estarían conectadas, formando una sola subpoblación (fig. 1).
Si bien gran parte de las amenazas han sido identificadas (Corbalán et al., 2010, 2023), la mayoría no han sido cuantificadas. Con base en las observaciones realizadas durante los muestreos en este estudio, se sistematizaron las amenazas para cada arroyo y cada subpoblación en el valle Pehuenche (tablas 2, 3). Como resultado de la priorización, las subpoblaciones del Límite y Rial Rojas se identifican como prioritarias para iniciar acciones concretas de conservación. Es posible que ambas sean fragmentos de una población mayor y que hayan estado unidas cuando el arroyo intermedio no estaba seco. En la del Límite se podría iniciar la mitigación de la depredación por salmónidos y reducir los efectos de la ruta. Es necesario continuar con el seguimiento y cuantificación de las amenazas, considerar las aún no evaluadas, tales como las que pueden estar generando las actividades turísticas, el vertido de sal en la ruta, crecidas extraordinarias y otras potenciales como el tendido de líneas de alta tensión, planificada desde el Maule (Chile) al río Diamante (San Rafael, Argentina).
El avance logrado en 20 años de estudios sobre la biología y conservación de este anfibio endémico de Argentina y Chile nos brinda herramientas para sostener acciones concretas de manejo que alivien alguna de las amenazas y que aseguren la viabilidad de las poblaciones a largo plazo. Se espera que un programa de monitoreo anual permita avanzar en el conocimiento de las tendencias poblacionales a largo plazo. Las subpoblaciones del valle Pehuenche son las mejor conocidas hasta el momento y han sufrido disminuciones, fragmentaciones y también extinciones locales. La priorización de su estado de conservación brinda herramientas para implementar acciones necesarias a corto plazo.
Agradecimientos
Por la colaboración en el trabajo de campo, se agradece a estudiantes de la Tecnicatura en Conservación de la Naturaleza Sede Malargüe del IEF Núm. 9-016, especialmente a Julián Rodríguez, Armando Barros, Francisco Jofré y Pablo Lucero; también a estudiantes voluntarios y Graciela Ríos. A Karen Olate por la elaboración del mapa. Y a un revisor anónimo por sus aportes a la primera versión del manuscrito y a los dos revisores que realizaron aportes importantes para mejorar el manuscrito. La investigación se logró gracias al financiamiento del proyecto SIIP 06/M003 Resol. 3978/2022 UNCUYO y aportes personales de investigadoras y estudiantes.
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Priority areas for conservation based on endemic vascular plant species and their biocultural attributes: a case study in Sinaloa, Mexico
C. Rocío Álamo-Herrera a, María Clara Arteaga a, *, Rafael Bello-Bedoy b
a Instituto Politécnico Nacional, Centro Interdisciplinario de Investigación para el Desarrollo Integral Regional, Unidad Durango, Sigma No. 119, Fracc. 20 de Noviembre II, 34234 Victoria de Durango, Durango, Mexico
b Universidad de Guadalajara, Centro Universitario de Ciencias Biológicas y Agropecuarias, Cátedras Conahcyt-Universidad de Guadalajara, Camino Ramón Padilla Sánchez No. 2100, 45200 Zapopan, Jalisco, Mexico
*Corresponding author: d1j17kk@hotmail.com (J.F. Pío-León)
Received: 20 February 2024; accepted: 02 July 2024
Abstract
Endemic vascular plants are one of the main biodiversity indicators used to propose priority conservation areas. The richness of endemic species and corrected and weighted endemism are the most frequently used criteria, while anthropogenic or biocultural factors such as ethnobotanical value or ecological vulnerability are seldom considered. This work proposes priority conservation areas for Sinaloa, Mexico, considering the richness of its endemic species, corrected and weighted endemism, as well as ethnobotanical value, protection status, and the Priority Conservation Index (PCI). The analysis was performed in a 19 × 19 km grid and included 247 records of 78 species. The areas proposed when considering only the richness of endemic species and the weighted endemism coincided with previously known areas of high biodiversity in the state, which are areas of high collection effort and low anthropogenic impact. When considering the ethnobotanical value and protection status, the areas identified included those with greater anthropogenic impact, which contained species of biocultural and economic importance. When the PCI was used, both of these types of regions were identified. We therefore recommend this index as a better indicator to select priority areas.
Keywords: Conservation index; Ebenopsis caesalpinioides; Ethnobotanical value; Protected Natural Areas; Priority species; Stenocereus martinezii
© 2024 Universidad Nacional Autónoma de México, Instituto de Biología. Este es un artículo Open Access bajo la licencia CC BY-NC-ND
(http://creativecommons.org/licenses/by-nc-nd/4.0/).
Áreas prioritarias para la conservación con base en especies de plantas vasculares endémicas y sus atributos bioculturales: un estudio de caso en Sinaloa, México
Resumen
Las plantas vasculares endémicas son uno de los principales indicadores empleados para proponer áreas prioritarias de conservación. La riqueza de especies endémicas y el endemismo ponderado y corregido son frecuentemente incluidos en los análisis, mientras que aspectos antropogénicos o bioculturales como el valor etnobotánico o la vulnerabilidad ecológica son poco considerados. Este trabajo propone áreas prioritarias de conservación para Sinaloa, México, considerando su riqueza de especies endémicas, endemismo ponderado y corregido, así como el valor etnobotánico, estatus de protección e índice prioritario de conservación (IPC). El análisis se realizó en cuadrículas de 19 × 19 km e incluyó 274 registros de 78 especies. Las áreas resultantes, considerando únicamente la riqueza de especies y el endemismo ponderado, coinciden con áreas previamente conocidas por su alta biodiversidad en el estado, mismas que poseen altos esfuerzos de colectas y bajos impactos antropogénicos. Por el contrario, cuando se consideró el valor etnobotánico y el estatus de protección, las áreas prioritarias incluyen zonas con alto impacto antropogénico, pero con presencia de especies con importancia biocultural y valor económico. Empleando el IPC se identificaron ambos tipos de regiones; en consecuencia, recomendamos este índice como un mejor indicador para seleccionar áreas prioritarias.
Palabras clave: Índice de conservación; Ebenopsis caesalpinioides; Valor etnobotánico; Áreas naturales protegidas; Especies prioritarias; Stenocereus martinezii
Introduction
Plants are essential organisms for maintaining the equilibrium of ecosystems and life on Earth. They provide the vast majority of the ecosystem and subsistence services that humans need to survive, including food, medicine, shelter, oxygen, carbon capture, and soil retention. Caring for plants is therefore an act of self-preservation (Raven, 2018). However, over 50% of the terrestrial vegetation on Earth is severely or moderately altered (Bradshaw et al., 2021).
Mexico is the country with the third to fifth highest plant richness, with more than 23,000 species, half of which are endemic (Conabio, 2023a; Villaseñor & Meave, 2022). However, despite 12% of Mexican territory being decreed as Protected Natural Area, it is estimated that between 37 and 50% of the nation’s land area has been impacted by human activities and that the majority of well-conserved areas are located in desert, semi-desert, and high mountain areas that are difficult to access (González-Abraham et al., 2015; Mora, 2019). Two of the largest and most biodiverse ecosystems in the country —dry forest and temperate forest— have suffered total degradation of 37 and 26% of their cover, respectively (Conabio, 2023b; Ulloa-Ulloa et al., 2017). The main causes of this deforestation have been agriculture and infrastructure development, both in Mexico specifically and worldwide (González-Abraham et al., 2015; Laso-Bayas et al., 2022).
One of the main analytical approaches used to propose priority conservation areas is grid analysis, which identifies centers of high biodiversity (“hotspots”) using criteria such as species richness, richness of endemic species, weighted endemism (WE), presence of threatened species, diversity of specific taxa (families or genera), or phylogenetic richness (Gutiérrez-Rodríguez et al., 2022; Maassoumi & Ashouri, 2022; Mehta et al., 2023; Murillo-Pérez et al., 2022; Qin et al., 2022; Sosa & De-Nova, 2012; Vargas-Amado et al., 2020; Villaseñor et al., 2022). The richness of endemic species in particular has the advantage of using a more precise (though smaller) database than the other aforementioned criteria for grid analysis to indicate conservation priority areas.
On the other hand, other indices can be used to propose conservation priority species based on their ethnobotanical or biocultural value, or the degree of threat they face due to use (e.g., Value of Use, Frequency of Use, Conservation Index) (De Lucena et al., 2013; Dhar et al., 2000; Mehta et al., 2023; Pío-León et al., 2023). However, these indices are not usually included in grid richness analyses to select priority conservation areas. These indices weight each species’ value based on its conservation priority, such that a priority conservation area would be determined not just by the total number of species or endemism, but also by their qualities.
Pío-León et al. (2023) compiled a list of the vascular plant species of Sinaloa and proposed some priority conservation areas based on the presence of 2 or more endemic species. In addition, the authors proposed a Priority Conservation Index (PCI) for each species based on its ethnobotanical value and ecological vulnerability, considering characteristics such as its distribution, habitat, and anthropogenic threats. In this index, species with high ethnobotanical value, slow growth (arboreal habit), threatened habitat (near to agricultural zones), and small distribution area (1 or a few known localities), have higher priority than those with no known ethnobotanical value, rapid growth (herbs), inaccessible habitat (cliffs or steep slopes), and wide distribution. The PCI was calculated with the formula:
PCI= D + H + Fv + Am + VE + Vc
where D is distribution; H, habitat; Fv, life form or habit (Spanish abbreviation); Am, degree of threat to their populations; VE, ethnobotanical value, and Vc, commercial value. However, that work did not perform a grid richness analysis to incorporate the values of these indices with traditional algorithms such as WE.
In the present work, we propose priority conservation areas in Sinaloa considering 3 types of algorithms: 1) richness of endemic species, WE, and corrected weighted endemism (CWE); 2) ethnobotanical value, protection status (NOM-059-SEMARNAT-2010 or IUCN) and PCI, and 3) the combination of 1) and 2). We hypothesized that incorporating those anthropogenic and biocultural attributes would modify the priority conservation areas selected since they will not necessarily correspond to the areas of the highest species richness.
Materials and methods
Sinaloa is located in northwestern Mexico, bordered on the east by the Sierra Madre Occidental (SMO) and on the west by the Pacific Ocean. According to Wiken et al. (2011), the main level III ecoregions that compose it are: 1) Sinaloa and Sonora Hills and Canyons with Xeric Shrub and Low Tropical Deciduous Forest (SS-TDF) (50%, 27,568 km2), which is located in the low parts of the SMO; 2) Sinaloa Coastal Plain with Low Tropical Thorn Forest and Wetland (S-TF) (29%, 15,612 km2), located in the lowlands near the coast, from the south-central portion northward; and 3) SMO with Conifer, Oak, and Mixed Forests (PQF) (15.78%, 8,681 km2) in the high parts of the western slope of the SMO (Fig. 1). It has been estimated that 4,000 species of vascular plants, nearly 80 of them endemic, occur in Sinaloa (Pío-León et al., 2023; Vega-Aviña et al., 2021). However, a large part of the coastal territory has been converted to agricultural land (~ 28,000 km2) (INEGI, 2023), resulting in severely fragmented habitats.
The database was based on the file generated by Pío-León et al. (2023) with some updates (Table 1). We incorporated recently described species (until October 2023) and removed species and collections that lacked reliable geographic coordinates. In addition, we prepared a matrix of weighted values considering the ethnobotanical value (E; 1 = documented use, 0 = no documented use), inclusion in a risk category (R) by the NOM-059-SEMARNAT-2010 (Semarnat, 2019) or IUCN (2023) (1 = included in at least 1 category, 0 = not included), and the value of the Conservation Priority Index (PCI), using the values reported by Pío-León et al. (2023) (Table 1). For the PCI, we assigned values according to their quartile position: 4 (upper quartile), 3 (second quartile), 2 (third quartile), and 1 (lower quartile). From these data, we formed 3 analysis groups: 1) biocultural value (E+R; 0 to 2), 2) PCI value (1 to 4), and 3) PCI + R (1 to 5).
Figure 1. Sinaloa state, Mexico, its main ecoregions level III (Wiken et al., 2011), and the regions of endemism according to Pío-León et al. (2023). Ecoregions: PQF = Conifer, Oak, and Mixed Forests of the Sierra Madre Occidental; SD = Sonoran Desert; SS-TDF = Sinaloa and Sonora Hills and Canyons with Xeric Shrub and Low Tropical Deciduous Forest; S-TF = Sinaloa Coastal Plain with Low Tropical Thorn Forest and Wetlands. Regions of endemism: 1 = Maviri-Topolobampo, 2 = Surutato region, 3 = Cerro Tecomate, 4 = Cerro Colorado, 5 = Sierra Tacuichamona, 6 = Meseta de Cacaxtla, 7 = Sierra de Concordia.
Table 1
List of endemic species of Sinaloa considered for this study and their scores by attributes. E = Ethnobotanical value; R = species with conservation status by the NOM-059-SEMARNAT-2010 or the IUCN (risk); PCI = Priority Conservation Index according to their quartile position.
| Especies | E | R | PCI | E+R | PCI+R |
| Acourtia gentryi L. Cabrera | 0 | 0 | 2 | 0 | 2 |
| Acourtia sinaloana B.L. Turner | 0 | 0 | 1 | 0 | 1 |
| Ageratina concordiana B.L. Turner | 0 | 0 | 2 | 0 | 2 |
| Albizia ortegae Britton & Rose | 0 | 0 | 3 | 0 | 3 |
| Aloysia nahuire A.H. Gentry & Moldenke | 1 | 0 | 4 | 1 | 5 |
| Anemia brandegeei Davenp. | 0 | 0 | 1 | 0 | 1 |
| Arachnothryx sinaloae Borhidi | 0 | 0 | 2 | 0 | 2 |
| Bastardiastrum tarasoides Fryxell | 0 | 0 | 2 | 0 | 2 |
| Bastardiastrum wissaduloides (Baker f.) Bates | 0 | 0 | 1 | 0 | 1 |
| Bletia santosii H. Ávila, J.G. González & Art. Castro | 0 | 0 | 2 | 0 | 2 |
| Bourreria franciscoi Pío-León & Vega | 0 | 0 | 3 | 0 | 3 |
| Bourreria ritovegana Pio-León, M.G. Chávez & L.O. Alvarado | 0 | 0 | 3 | 0 | 3 |
| Bouvardia sinaloae Borhidi & E. Martínez | 0 | 0 | 2 | 0 | 2 |
| Calliandra estebanensis H.M. Hern. | 0 | 0 | 2 | 0 | 2 |
| Carlowrightia fuertensis T.F. Daniel | 0 | 0 | 3 | 0 | 3 |
| Castilleja racemosa (Breedlove & Heckard) T.I. Chuang & Heckard | 0 | 0 | 3 | 0 | 3 |
| Chrysactinia lehtoae D.J. Keil | 0 | 0 | 2 | 0 | 2 |
| Cnidoscolus sinaloensis Breckon ex Fern.Casas | 0 | 1 | 3 | 1 | 4 |
| Cochemiea thomasii García-Mor., Rodr. González, J. García-Jim. & Iamonico | 0 | 0 | 2 | 0 | 2 |
| Coutaportla helgae Pío-León, Torr.-Montúfar & H. Ávila | 0 | 0 | 1 | 0 | 1 |
| Coutaportla lorenceana Torr.-Montúfar, H. Ochot. & Art.Castro | 0 | 0 | 2 | 0 | 2 |
| Croton ortegae Standl. | 0 | 0 | 3 | 0 | 3 |
| Ctenodon rosei Morton | 0 | 0 | 3 | 0 | 3 |
| Cuphea delicatula Brandegee | 0 | 0 | 2 | 0 | 2 |
| Cyclanthera monticola Gentry | 0 | 0 | 2 | 0 | 2 |
| Dioscorea sinaloensis O. Téllez | 0 | 0 | 2 | 0 | 2 |
| Dryopetalon breedlovei (Rollins) Al-Shehbaz | 0 | 0 | 1 | 0 | 1 |
| Ebenopsis caesalpinioides (Standl.) Britton & Rose | 1 | 1 | 4 | 2 | 5 |
| Echeveria coppii Moran ex Gideon F.Sm. & Bischofberger | 0 | 0 | 2 | 0 | 2 |
| Echeveria juliana Reyes, González-Zorzano & Kristen | 0 | 0 | 1 | 0 | 1 |
| Echeveria kimnachii J. Meyrán & R. Vega | 0 | 0 | 1 | 0 | 1 |
| Epidendrum petacaense Hágsater, J. Duarte & Pío-León | 0 | 0 | 2 | 0 | 2 |
| Eryngiophyllum rosei Greenm. | 0 | 0 | 2 | 0 | 2 |
| Frangula surotatensis (Gentry) A. Pool | 0 | 0 | 2 | 0 | 2 |
| Graptopetalum sinaloensis Vega | 0 | 0 | 1 | 0 | 1 |
| Guardiola stenodonta S.F. Blake | 0 | 0 | 1 | 0 | 1 |
| Helicteres vegae Cristóbal | 0 | 0 | 3 | 0 | 3 |
| Heliopsis sinaloensis B.L. Turner | 0 | 0 | 2 | 0 | 2 |
| Table 1. Continued | |||||
| Especies | E | R | PCI | E+R | PCI+R |
| Hofmeisteria sinaloensis Gentry | 0 | 0 | 1 | 0 | 1 |
| Indigofera sinaloensis M. Sousa & Cruz Durán | 0 | 0 | 2 | 0 | 2 |
| Ipomopsis monticola J.M. Porter & L.A. Johnson | 0 | 0 | 2 | 0 | 2 |
| Iresine arenaria Standl. | 0 | 0 | 1 | 0 | 1 |
| Koanophyllum concordianum B.L. Turner | 0 | 0 | 2 | 0 | 2 |
| Lasianthaea gentryi B.L. Turner | 0 | 0 | 2 | 0 | 2 |
| Lasianthaea ritovegana B.L. Turner | 0 | 0 | 1 | 0 | 1 |
| Licania mexicana Lundell | 0 | 0 | 2 | 0 | 2 |
| Lobelia macrocentron (Benth.) T.J. Ayers | 0 | 0 | 2 | 0 | 2 |
| Lopezia conjugens Brandegee | 0 | 0 | 1 | 0 | 1 |
| Lopezia sinaloensis Munz | 0 | 0 | 1 | 0 | 1 |
| Lupinus gentryanus C.P. Sm. | 0 | 0 | 2 | 0 | 2 |
| Lupinus howard-scottii C.P. Sm. | 0 | 0 | 2 | 0 | 2 |
| Lupinus sinaloensis C.P. Sm. | 0 | 0 | 2 | 0 | 2 |
| Mariosousa gentryi Seigler & Ebinger | 0 | 0 | 3 | 0 | 3 |
| Mimosa coelocarpa B.L. Rob. | 0 | 0 | 3 | 0 | 3 |
| Mitracarpus aristatus Borhidi & Lozada-Pérez | 0 | 0 | 2 | 0 | 2 |
| Molinadendron sinaloense (Standl. & Gentry) P.K. Endress | 0 | 1 | 3 | 1 | 4 |
| Pavonia gentryi Fryxell | 0 | 0 | 2 | 0 | 2 |
| Peniocereus papillosus (Britton & Rose) U. Guzmán | 0 | 0 | 2 | 0 | 2 |
| Periptera trichostemon Bullock | 0 | 0 | 2 | 0 | 2 |
| Perityle canescens Everly | 0 | 0 | 1 | 0 | 1 |
| Perityle grandifolia Brandegee | 0 | 0 | 1 | 0 | 1 |
| Perityle stevensii B.L. Turner | 0 | 0 | 1 | 0 | 1 |
| Physalis vestita Waterf. | 0 | 0 | 3 | 0 | 3 |
| Pitcairnia monticola Brandegee | 0 | 0 | 1 | 0 | 1 |
| Polygala polyedra Brandegee | 0 | 0 | 1 | 0 | 1 |
| Psacalium quercifolium H.Rob. & Brettell | 0 | 0 | 2 | 0 | 2 |
| Salvia beltraniorum J.G.González, Pío-León & Art.Castro | 0 | 0 | 2 | 0 | 2 |
| Salvia trichostephana Epling | 0 | 0 | 2 | 0 | 2 |
| Sedum copalense Kimnach | 0 | 0 | 1 | 0 | 1 |
| Stenocereus martinezii (J.G. Ortega) Bravo | 1 | 1 | 4 | 2 | 5 |
| Stevia concordiana B.L. Turner | 0 | 0 | 2 | 0 | 2 |
| Sysyrinchium jacquelineanum Art.Castro, H. Ávila & J.G. González | 0 | 0 | 1 | 0 | 1 |
| Tibouchina thulia Todzia | 0 | 0 | 1 | 0 | 1 |
| Tillandsia mazatlanensis Rauh | 0 | 0 | 2 | 0 | 2 |
| Tillandsia occulta H. Luther | 0 | 0 | 2 | 0 | 2 |
| Verbesina microcarpa S.F. Blake | 0 | 0 | 2 | 0 | 2 |
| Verbesina ortegae S.F. Blake | 0 | 0 | 2 | 0 | 2 |
| Verbesina sinaloensis B.L. Turner | 0 | 0 | 2 | 0 | 2 |
Richness of endemism (SR), weighted endemism (WE), and corrected weighted endemism (CWE). The richness of endemic species was quantified in 19 × 19 km cells (361 km2), dividing Sinaloa into 195 cells. The cell size used was determined according to the criterion of Oyala (2020). Endemic species richness was quantified as the total number of endemic species whose distribution includes the cell. Endemism was evaluated using the WE and CWE indices. The WE score for each cell was obtained by summing, for each species present in the cell, the inverse of the number of cells in which the species occurs; thus, a high WE value indicates cells that contain more species with restricted distributions (i.e., that are found in few other cells), while low WE values indicate cells that mostly contain widely distributed species (i.e., species that are also present in other cells). The CWE is similar, but additionally corrects for potential biases due to differences in overall richness by dividing the value of the WE by the number of species present in the cell (Laffan & Crisp, 2003). The 3 parameters (SR, WE, and CWE) were estimated in the program Biodiverse v.2.0 (Laffan et al., 2010). Geoprocessing of the data was performed in QGIS 3.4.8 (QGIS.org, 2019).
Endemism weighted by biocultural attributes and PCI. In addition to SR, WE, and CWE analysis, endemism weighted by biocultural attributes was evaluated using 2 sets of attribute/parameter combinations, each resulting in 3 maps, 9 in total (Fig. 2). The first set included the species richness plus the biocultural values, resulting in the following 3 combinations: species richness plus biocultural value (SR+E+R), species richness plus PCI (SR+PCI), and PCI plus the risk category (SR+PCI+R). The second set did not consider species richness, resulting in the combinations of biocultural value (E+R), PCI, and PCI+R. For this second set of analyses, only species that fulfilled the relevant criteria were included (e.g., the E+R combination included only species that had ethnobotanical value and are included in a risk category). As such, in the first set of maps, a priority conservation area depended by the number of species present and their qualities (e.g., species with ethnobotanical value or species with protected status), while in the second only the species’ qualities were considered.
The final priority conservation areas were based on the consensus map of the 9 different endemism maps. The consensus areas took into account only the cells that had the highest possible value of the relevant variables in at least 1 of the 9 previously generated endemism maps. The consensus values were obtained by summing the number of times each cell had the highest possible value in each of the endemism maps, such that the highest possible consensus value was theoretically 9 (the cell had the highest possible value in all maps), and the minimum value was 1 (maximal value in only 1 map). The consensus map was also overlayed with Protected Natural Areas and Priority Terrestrial Regions, land use, and bioclimatic corridors.
Results
Occurrence, conservation (risk) status, and ethnobotanical uses of the endemic species of Sinaloa. The database contained 247 records of 78 species, 30 families, and 61 genera. For 48 of the genera (78.7%), only 1 species of the genus was present. The majority of the records were distributed in the central to the southern region of the state, near the coast, in the Meseta de Cacaxtla Natural Protected Area and the area between the former and Sierra de Tacuichamona, as well as in the Concordia and Surutato mountains of the SMO (Fig. 3; regions 6, 5, and 2, in Figure 1). The 2 level III ecoregions best represented were SS-TDF (166 records/ 40 species) and the PQF (53/ 37), followed by the S-TF (17/ 8) (Fig. 3a). Sixty-nine percent of the records fell outside of the polygons of Protected Natural Areas or Priority Terrestrial Conservation Regions (Fig. 3B). Sixty-eight percent of the species (53) were known from a single locality (either a single collection or collections from locations that are very close to each other).
Figure 2. Flowchart of the endemism analysis to select priority areas in Sinaloa, Mexico.
Only 4 species (Cnidoscolus sinaloensis, Ebenopsis caesalpinioides, Molinadendron sinaloense, and Stenocereus martinezii) of the 78 analyzed are found in some risk category (Table 1). All 4 are considered endangered (EN) by the IUCN, while only Stenocereus martinezii is included in NOM-059-SEMARNAT-2010, under the category of special protection (Pr). Only 3 species have well-documented ethnobotanical uses: Aloysia nahuire (aromatic and medicinal tea), Ebenopsis caesalpinioides (edible seeds, occasional commercial value), and Stenocereus martinezii (edible fruits, commercial value). One additional species, Lupinus gentryianus, was noted in the type collection to be used as an anti-parasitic for livestock; however, this plant is only known from that locality, and this use has not since been confirmed, so it was not considered.
The different patterns of endemism are shown in Figure 4. The overall richness of endemism (Fig. 4A) showed 2 main areas —1 in the northern part of the Sierra de Concordia (region 7, Fig. 1) and the other in the western part of the Sierra de Tacuichamona (region 5, Fig. 1)— as well as 3 secondary areas located in the Sierra de Surutato (region 2, Fig. 1), Cerro Colorado (region 4, Fig. 1), and the southern part of the Sierra de Concordia. The WE (Fig. 4B) showed a similar pattern in richness but with an increase in the priority levels of the Sierra de Surutato and a decrease by 1 level for the Tacuichamona and Cerro Colorado. The CWE (Fig. 4C) showed several priority areas more scattered across the state than the WE, mainly in the SMO, corresponding to the majority of the species known from a single locality; however, compared with WE, there was a greater concentration of high-priority cells toward the northern part of the state, near southern Sonora, in the area around the Sierra de Barobampo and Hills of Topolobampo (region 1, Fig. 1).
Figure 3. Records of endemic species in Sinaloa overlayed onto: level III ecoregions (Wiken et al., 2011; definitions in Figure 1) (A) and Protected Natural Areas (PNA) and Priority Terrestrial Regions (B). Categories of Protected Natural Areas: PNAS = state; PNAM = municipal; PNAF = federal; PTR = Priority Terrestrial Regions.
Figure 4. Endemism areas of vascular plants in Sinaloa, Mexico, according to the calculated index values (A-I): SR = endemic species richness; WE = weighted endemism; CWE = corrected weighted endemism; E = ethnobotanical value; R = species with protection status; PCI = Priority Conservation Index.
Figure 5. Consensus priority conservation areas (PAC) in the state of Sinaloa (A-D). Consensus map (A) superimposed to: Protected Natural Areas/Priority Terrestrial Regions (B), land use (C), and bioclimatic corridors (D). Categories of Protected Natural Areas: PNAS = state; PNAM = municipal; PNAF = federal; PTR = Priority Terrestrial Regions.
The addition of the ethnobotanical attributes to the protection status and richness of endemic species (SR+E+R) (Fig. 4D) showed an increase in the values for the areas from the Meseta de Cacaxtla (region 6, Fig. 1) to Sierra de Tacuichamona, but a decrease in the zones of the SMO. Adding the Priority Conservation Index to the richness (SR+PCI) (Fig. 4E) showed an increase and homogenization of the priority in all of the aforementioned regions, while adding protection status (SR+PCI+R) (Fig. 3F) did not significantly modify the areas of importance.
Finally, when considering only the ethnobotanical value plus the protection status (E+R), without considering species richness (i.e., eliminating the species that did not have those attributes), the zone of highest priority was concentrated nearly exclusively in the southern part of the state, within and adjacent to the Meseta de Cacaxtla (Fig. 3G). When considering PCI only or PCI plus risk category, there was again a homogenization of the high priority for the 2 mountainous areas (Surutato and Concordia), Meseta de Cacaxtla, Tacuichamona, and surrounding areas (Fig. 4H, I).
The priority conservation areas, as defined by the consensus among the 9 maps analyzed, were composed of 7 polygons grouped into 3 categories (Fig. 5): 4 cells with a value of 6 (of the maximum possible score of 9) in the northern part of the Sierra de Concordia, northwestern part of Sierra Surutato, Meseta de Cacaxtla, and Sierra de Tacuichamona; 1 with a value of 4 in the area between the Meseta de Cacaxtla and Tacuichamona; and 2 with a value of 3 in the southern part of the Sierra de Concordia and southeastern part of the Sierra de Surutato (Fig. 5A). However, since the 3 areas with a value of 3 or 4 were contiguous with areas with a value of 6, 4 priority conservation areas were proposed: Sierra de Surutato (Fig. 45-a), Sierra de Tacuichamona (Fig. 5A-b), Meseta de Cacaxtla (Fig. 5A-c), and Sierra de Concordia (Fig. 5A-d).
Superimposing the consensus map with the map of existing Protected Natural Areas (Fig. 5B) showed that these 4 consensus areas fall partially within protected areas: 1 federal (Área de Protección de Flora y Fauna Meseta de Cacaxtla, Fig. 5B-c), 1 state (Sierra de Tacuichamona, Fig. 5B-b), and 2 municipal (Reserva Chara Pinta in the Sierra de Concordia and Reserva de Surutato, Fig. 5B-d, B-a, respectively). The Sierra de Concordia also includes part of the terrestrial priority region Río Presidio. Regarding land use, the 2 consensus areas in the SMO were found in mixed pine-oak forest with low impact of agricultural activity (Fig. 5C-a, C-d), while the other 2, located in the Sinaloa and Sonora Hills and Canyons with Xeric Shrub and Low Tropical Deciduous Forest ecoregion, present moderate to high impact from irrigated and rainfed agriculture (Fig. 5C-b, C-c). When considering biological corridors, only the priority area in the Meseta de Cacaxtla overlapped with a bioclimatic corridor.
Discussion
The analyses of richness of endemic species and WE showed higher conservation priority in areas that were previously identified as having high endemism (Pío-León et al., 2023), low anthropogenic impact from agriculture, and which have also historically been subject to concentrated collection efforts (Sierra de Surutato and Sierra de Concordia) (Ávila-González et al., 2019; Gentry 1946; Vega-Aviña et al., 2021). On the other hand, the regions defined based on CWE reflected a high number of species known from a single locality, which could indicate the presence of small islands of endemism in the state or low collection effort. In contrast, the inclusion of the ethnobotanical criteria and protection status (E+R) shows a different pattern from species richness, concentrating high priority scored in an area of transition between the coastal plain of Sinaloa and the hills of Sinaloa and Sonora, near the coast in the center-south of the state. These regions correspond to the transition and ecotone between low tropical deciduous forest and thorn forest, which are strongly impacted by anthropogenic activities (irrigated and rainfed agriculture), suggesting that the species with the highest ethnobotanical importance and with protected status (IUCN or NOM-053-SEMARNAT-2010) are found near human activities that require stronger conservation attention than those located in the high parts of the SMO, where the threats are less severe.
The priority conservation areas indicated by the consensus map (Fig. 5) include the regions with the highest richness of endemic species plus the areas with the highest number of species with biocultural importance. These consensus areas are practically the same as those that were assigned the highest priority values when considering only the Priority Conservation Index (PCI) for each species; as such, this index was the most robust single indicator for selecting priority conservation areas. This index combines ethnobotanical parameters such as species’ uses and economic value with ecological parameters such as their distribution, habit, and habitat. Thus, it covers a broad range of criteria that are useful for defining priority species or areas for conservation.
All the priority conservation areas defined by the consensus map (Fig. 5) except 1 included part of a Protected Area polygon, although only 1 was under federal jurisdiction (Meseta de Cacaxtla). The only cell that did not overlap with a Protected Natural Area was adjacent to the Meseta de Cacaxtla, and it was the cell with the largest area of agriculture. This area is important because it contains the 2 species with the highest ethnobotanical value (Ebenopsis caesalpinioides and Stenocereus martinezii), which are also found in a risk category according to the IUCN and NOM-053-SEMARNAT-2010. This area therefore urgently requires conservation and restoration activities, especially for E. caesalpinioides, whose distribution is limited to the area surrounding this cell (Pío-León et al., 2023). Specifically, we recommend avoiding the conversion from rainfed agriculture to technified irrigated agricultural activities, since these are generally more aggressive toward native vegetation. This area is also important because it is located at the transition between lowland deciduous forest and thorn forest of Sinaloa, which could reflect high endemism, in addition to potentially serving as part of the bioclimatic corridor connecting the 2 most important terrestrial ANPs in the state, Meseta de Cacaxtla (federal) and Sierra Tacuichamona (state).
In the present study, the incorporation of the species’ biocultural parameters modified the priority areas for conservation compared to the areas selected when considering only the richness of endemic species, weighted endemism, or corrected weighted endemism. Specifically, the richness analysis identified priority areas in the mountainous and high-diversity regions of Sinaloa, while the ethnobotanical and ecological factors incorporated zones near the coast that have higher anthropogenic impact. The Conservation Priority Index identified all of these priority regions; for this reason, we propose it as a complete and robust index for identifying priority conservation areas. At the state level, we recommend that conservation and restoration actions be implemented in the area of transition between the low tropical deciduous forest and thorn forest. This area simultaneously presents the highest impact of anthropogenic activities and harbors the most important Sinaloa endemic species in terms of biocultural value and protection status —the “pitaya de Sinaloa” (Stenocereus martinezii) and the “guampinola” or “frutilla” (Ebenopsis caesalpinioides). This area should be considered a priority for both conservation and restoration, which would not have been identified as a priority if only the richness of endemism or CWE had been analyzed.
Acknowledgements
The first author is grateful to the Consejo Nacional de Humanidades, Ciencia y Tecnología (Conahcyt) for the grant awarded as part of the Estancias Posdoctorales por México program (I1200/320/2022). We also thank Jorge David López Pérez for his suggestions on data analysis, and the two anonymous reviewers for their comments and suggestions that improved our manuscript.
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vel-iii
Mitogenome of the Golden Eagle (Aquila chrysaetos) in northwestern Baja California, Mexico: phylogenetic relationships and genetic variation
Francisco J. García-De León a, Gorgonio Ruiz-Campos b, *, Jesús Roberto Oyervides-Figueroa c, Gonzalo De León-Girón b, Carlos Alberto Flores-López b, Dante Magdaleno-Moncayo d, Alicia Abadía-Cardoso e
a Centro de Investigaciones Biológicas del Noroeste, S.C., Laboratorio de Genética para la Conservación, Av. Instituto Politécnico Nacional 195, Playa Palo de Santa Rita Sur, 23096 La Paz, Baja California Sur, Mexico
b Universidad Autónoma de Baja California, Facultad de Ciencias, Carretera Ensenada-Tijuana Km. 103 s/n, 22860 Ensenada, Baja California, Mexico
c Centro de Investigación Científica y de Educación Superior de Ensenada, Departamento de Acuicultura, Carretera Tijuana -Ensenada 3918, Zona Playitas, 22860 Ensenada, B.C., Mexico
d Universidad Autónoma de Baja California, Facultad de Ingeniería, Arquitectura y Diseño, Carretera Ensenada-Tijuana Km. 103 s/n, 22860 Ensenada, Baja California, Mexico
e Universidad Autónoma de Baja California, Facultad de Ciencias Marinas, Carretera Ensenada-Tijuana Km. 103 s/n, 22860 Ensenada, Baja California, Mexico
*Corresponding author: gruiz@uabc.edu.mx (G. Ruiz-Campos)
Received: 20 August 2024; accepted: 20 February 2025
Abstract
We assembled and annotated the mitochondrial genome of golden eagles from the northwest of Baja California, Mexico, using reference and de novo strategies to analyze the synteny of mitochondrial genes, the phylogenetic relationships, and genetic variation of mitochondrial DNA. The length of the Golden Eagle mitogenome was 17,472 bp (base pairs) with a base composition of A (29.8%), C (32.5%), G (14.0%), and T (23.6%). The mitogenome contains 13 genes coding for the protein complexes coxI II-III, Cytb, cATP 6 and 8, and Nicotinamide Adenine Dinucleotide (NADH 1-6); this arrangement is consistent with the general model of the mitogenome reported in other congeneric members. Mitogenomes of individuals from northwestern Baja California are unique and they differ from the mitogenome of southern California golden eagles in 3 traits: 1) molecule size is 140 bp larger than that previously reported, 2) the addition in the annotation of a region called pseudo control (ψRC), and 3) the annotation in 2 fractions of the coding region for the protein NADH dehydrogenase subunit 3 (ND3). The genetic diversity and phylogenetic analyses of the individual genes and mitogenome support a close genetic relatedness between the golden eagles from northwestern Baja California and southern California region.
Keywords: Mediterranean region; Mitochondrial lineages; Mitochondrial genome; Next generation sequencing; Non-model species
Mitogenoma de águila real (Aquila chrysaetos) en el noroeste de Baja California, México: relaciones filogenéticas y variación genética
Resumen
Ensamblamos y anotamos el genoma mitocondrial del águila real del noroeste de Baja California, México, utilizando estrategias de referencia y de novo para analizar la sintenia de genes mitocondriales, las relaciones filogenéticas y la variación genética del DNA mitocondrial. La longitud del mitogenoma fue 17,472 pb (pares de bases) con una composición de bases de A (29.8%), C (32.5%), G (14.0%) y T (23.6%). El mitogenoma contiene 13 genes codificantes de los complejos proteínicos coxI II-III, Cytb, cATP 6 y 8, y nicotinamida adenina dinucleótido (NADH 1-6); esto fue consistente con el modelo general del mitogenoma reportado en otros congéneres. Los mitogenomas de individuos del noroeste de Baja California son únicos y se diferencian del mitogenoma de individuos del sur de California en 3 rasgos: 1) el tamaño de molécula es 140 pb más grande que el reportado, 2) adición de la región llamada pseudocontrol (ψRC) y 3) anotación en 2 fracciones de la región codificante de la proteína NADH deshidrogenasa subunidad 3 (ND3). La diversidad genética y los análisis filogenéticos de los genes individuales y el mitogenoma respaldan una estrecha relación genética entre las águilas reales del noroeste de Baja California y la región del sur de California.
Palabras clave: Región mediterránea; Linajes mitocondriales; Genoma mitocondrial; Secuenciación de próxima generación; Especies no modelo
Introduction
The Golden Eagle, Aquila chrysaetos, is an accipitrid of boreal distribution that inhabits a variety of open and semi-open biotopes from sea level to 3,630 m in altitude in biomes as tundra, chaparral, temperate grassland, temperate deciduous forest, and coniferous forest (De León-Girón et al., 2016; Flesch et al., 2020; Kochert et al., 2002; Watson et al., 2011). This emblematic species is known to occur in Mexico in the arid and semiarid environments of the northern and central regions, including some locations as south as Oaxaca (Bolger et al., 2014; De León-Girón et al., 2016; Howell & Webb, 1995; Rodríguez-Estrella, 2002; Rodríguez-Estrella et al., 1991, 2020). Although the demography of the Golden Eagle has been determined with different genetic methods that confirm the structuring of its populations (Craig et al., 2016; Doyle et al., 2016), little is known about the genetic identity of populations of this eagle in its distribution range. Northwestern Baja California is considered the greatest nesting and conservation potential region for the Golden Eagle in Mexico (De León-Girón et al., 2016; Rodríguez-Estrella, 2002; Rodríguez-Estrella et al., 1991; Tracey et al., 2017), with vast areas of habitats still pristine for the population conservation of this species shared with the United States of America (Craig et al., 2016; De León-Girón et al., 2016, 2024; Doyle et al., 2014; Katzner et al., 2023).
Studies based on mitochondrial DNA (mtDNA) allow determining the current state of conservation of species, as well as the identification of significant evolutionary and management units with their demographic aspects (Moritz, 1994). With the advent of next-generation sequencing (NGS) technologies (especially on complete mitochondrial genomes), new studies with non-model species became more common. As was the case for various species of the family Accipitridae, including the Golden Eagle. Doyle et al. (2014) were the first to describe the nuclear and mitochondrial genome of the Golden Eagle from the central California region, in the USA.
In this study we focus on the synteny of mitochondrial genes, the phylogenetic relationships, and the mitochondrial genetic diversity among Golden Eagle individuals in northwestern Baja California, Mexico. Given the high migration potential of this species, individuals collected in northwestern Baja California are predicted to be phylogenetically related to the westernmost Golden Eagle population in the USA. Therefore, studies focused on examining the genetic variation at the mitochondrial level will allow better decision-making for management and conservation programs for the species in a binational spectrum. In the same way, our study will be a valuable reference for analyzing other Golden Eagle populations in Mexico and other regions of its distribution range.
Figure 1. Geographical location of the Golden Eagle voucher specimens examined for mitogenome in the northwestern Baja California, Mexico. Geographic coordinates for each individual can be found in Supplementary material T1. Map by Rafael Hernández Guzmán.
Materials and methods
Tissue samples of 6 Golden Eagle specimens from northwestern Baja California, Mexico were obtained. These specimens were found dead in different agriculture valleys of the Mediterranean region in northwestern Baja California, Mexico, between the municipalities of Rosarito and San Quintín, from 1995 to 2014, and deposited as vouchers in the Bird Collection of the Science Faculty, at the Autonomous University of Baja California (UABC), campus Ensenada (Fig. 1, Supplementary material: T1).
For the DNA extraction, tissue samples of the pads of the feet of 5 Golden Eagle specimens referred to here as Ach 1 to Ach 5 were obtained. DNA using the nucleic acid purification method by differential saline precipitation (Aljanabi & Martinez, 1997) was extracted. Both feathers and blood for the Ach 6 sample (Supplementary material: T1) were used and extracted DNA with the Qiagen Blood & Tissue kit. The quality and quantity of the extracted DNA in both cases was evaluated using Nanodrop, Qubit and agarose gel electrophoresis.
Extracted DNA (from 77 to 210 ng/µL per sample quantified in Nanodrop for Ach1-Ach5 individuals and 0.581 ng/µL for Ach6 quantified in Qubit) was purified with SpeedBeads magnetic beads (Thermo-Scientific, Waltham, MA, USA) at a 1.2:1 beads:DNA ratio and resuspended in 50 µl of TLE1X buffer. Libraries for shotgun sequencing were prepared, for which purified DNA from each individual was fragmented by sonication in a Bioruptor® (Diagenode, Liege, Belgium) using 2 rounds, each consisting of 5 cycles of 30 sec of sonication and 30 sec without sonication at the highest setting. DNA fragments were then prepared using the Kapa Biosystems® Hyper Prep Kit (KR0961–v4.15, Roche, Basel, Switzerland), with which end repair and A-tailing were performed, adapters were ligated, and PCR was performed with indexed primers (Glenn et al., 2020) for 14 cycles. Following amplification, fragment size selection was performed by double-dip with SpeedBeads magnetic beads (Thermo-Scientific, Waltham, MA, USA) that allow to preserve fragments between ~250 and ~700 base pairs (bp) for each of the samples, which are the appropriate insert size for the Illumina platform. Sequencing was performed paired end on 2 different platforms. We sequenced the Ach1-Ach5 samples on the Illumina MiSeq V.3 at the Georgia Genomics and Bioinformatics Core (GGBC) to generate 300-bp paired-end fragments, and we sequenced Ach6 on an Illumina HiSeq 4000 at the Oklahoma Medical Research Foundation Clinical Genomics Center to generate 150-bp paired-end fragments.
Bioinformatic analysis. The quality of the raw sequences was evaluated using the FastQC program (Andrews, 2010). Subsequently, the sequences using a standard treatment were filtered using the Trimmomatic software (Bolger et al., 2014) with 4 steps. In the first step, the sequencing adapters and over-represented sequences were removed by means of the ILLUMINACLIP function. The second step discarded sequences with a quality value below 30 QS with the AVGQUAL function. The third step removed fragments below 25 QS with the SLIDINGWINDOW function. The fourth step did an even stricter cleanup using MAXINFO. This step conserved sequences of at least 100 base pairs and was configured to preferentially conserve longer sequences with a value of 0.3. For the Ach 5 specimen, an extended and modified version of the standard treatment was used due to the low quality of sequences. This involved applying the SLIDINGWINDOW function, which modifies the quality value.
Assembly and annotation. Two strategies were followed to assemble the mitochondrial genomes: the reference based and the de novo strategy (Machado et al., 2015, 2018). Mapping against a reference mtDNA genome was performed using the Bowtie2 program version 2.3.4.2 with clean reads (Langmead & Salzberg, 2012). The reference mitochondrial genome was from the same species A. chrysaetos, with the GenBank accession number of KF905228.1. De novo assembly only for samples Ach 2 to Ach 6 was performed using the A5-miseq pipeline (Coil et al., 2015). Due to the large number of reads after the filter (Supplementary material: T2) sample Ach 1 was analyzed with the Velvet software version 1.1 (Zerbino & Birney, 2008). The mitogenome was annotated from the scaffolds obtained from the de novo assemblies. We determined scaffolds longer than 4 Kpb by applying a search with the BLAST tool (Basic Local Alignment Search Tool) of the NCBI (National Center for Biotechnology Information) website. The scaffolds were aligned to achieve greater coverage considering the length of the reference genome used previously. Afterwards, RNAweasel (http://megasun.bch.umontreal.ca/RNAweasel/) was implemented to the identification of tRNA’s (transfer RNA), rRNA’s (ribosomal RNA) and introns and, in turn, MFannot (http://megasun.bch.umontreal.ca/RNAweasel/) for the identification of proteins and open reading frames (Sieber et al., 2018).
In addition to the previous annotation, MITOS web server (Bernt et al., 2013) was used to assist in the annotation of de novo mitochondrial genomes, allowing gene names, tRNA and rRNA secondary structures, and codon usage to be obtained. Finally, a manual curation of the annotations was used to review the 6 existing reading frames with the UGENE software (Okonechnikov et al., 2012). The control region was identified based on 99% similarity with an A. chrysaetos partial control region sequence (EF459579.1). Genome Vx software (Conant and Wolfe, 2008) was used to map the Golden Eagle’s mitogenome, using individual Ach 6 to achieve this analysis.
In the annotation of the de novo assembly, the ND3 protein appeared divided into 2 fractions (see Results), a trait that was not reported by Doyle et al. (2014). Therefore, an experimental verification was required through PCR amplification and its subsequent Sanger sequencing. We designed primers from de novo assembly using the NCBI Primer-BLAST program: ND3Ach-1 (5’GCCTGATACTGGCACTTCGT3’ and 5’CCCTATCAATCTGACCCACCG3’) which generates a 716 bp fragment and ND3Ach-2 (5’CTTCTTCGTCGCTACAGGCT3’ and 5’CCTTCCACCGAACCCACTTAA3’) which generates a 774 bp fragment. Eight Golden Eagle samples from the Ornithological Collection of the Faculty of Sciences of the Autonomous University of Baja California were used for PCR amplification. These samples correspond to the 6 individuals used for sequencing and 2 extra samples that are not part of the mitogenome assembly study. The conditions of the PCR reaction were as follows: 6 min at 94°C, 35 cycles of denaturation, annealing, and extension at 94°C for 30 sec, 51°C for 30 sec and at 72°C for 1.5 min, respectively. A final elongation stage at 70°C for 7 min, and a final conservation stage at 15°C. We sent the amplified fragments for sequencing to the SeqXcel.inc Company in San Diego, California, and analyzed with the ABI PRISM® 3130xl Genetic Analyzer DNA kit.
Synteny and search for polymorphisms. A posteriori synteny analysis was performed between the Golden Eagle mitogenome reported by Doyle et al. (2014) and the newly assembled mitochondrial genome of this study. This analysis was carried out with the MAUVE software (Darling et al., 2004). The same reference mitogenome was used to call SNP’s and INDEL’s for the 6 genomes assembled in this work using the SAMtools program (Li, 2011). Filtering the call quality of the variants involved discarding those with a value less than 20 Qs and keeping those with at least 5X depth (Li, 2011). Each variant was manually evaluated, retaining SNPs and INDELs where at least half of the aligned reads showed the change and verified its existence in at least one other individual (Fridjonsson et al., 2011).
Phylogenetic relationships. Phylogenetic relationships were inferred using the individual gene dataset and the mitogenome. The individual gene fragments analyzed consisted of the ND2, ND3, coxI and Cytb recovered from Aquila chrysaetos individuals from GenBank (Supplementary material: T3) and those produced in this study. The mitogenome analysis for phylogenetic relationships consisted of 10 complete mitogenomes: 6 obtained in this study, 2 reference samples of Aquila chrysaetos (LR822062 from the United Kingdom and NC_024087.1 from California, USA), 1 reference sample classified as Aquila heliaca (NC_035806.1), but has been reported to actually consist of a A. chrsyasetos individual (Sangster & Lukesenburg, 2021) and 1 outgroup (Aquila nipalensis, GenBank accession number NC_045042.1). Both datasets were aligned using the MAFFT 7 algorithm (Katoh & Standley, 2013). Removing the non-conserved regions between the sequences of the multiple alignments was done using the online program Gblocks with the default parameters (Castresana, 2000). The nucleotide substitution model and the mutation rate were determined using JModelTest (Posada, 2008), considering the Bayesian information criterion (BIC).
Two phylogenetic reconstruction methods were used: Maximum Likelihood (ML) and Bayesian analysis. In IQ-TREE (Nguyen et al., 2015) the ML method (Felsenstein, 1981) was run with the DNA substitution model selected by ModelFinder (Kalyaanamoorthy et al., 2017) and 1,000 bootstrap pseudo replicates were performed (Hoang et al., 2017). Bayesian inference was used with MrBayes program (Ronquist et al., 2012); for this, 4 Markov Monte Carlo chains (MCMC) were implemented with a total of 10 million generations, and a sampling every 1,000 generations. To construct the consensus phylogenetic tree, a 25% of burn-in was applied. The support of the nodes was evaluated via the posterior probability values. The results of both approaches were visualized with FigTree 1.4 (Rambaut, 2009). For the mitogenome phylogenetic tree, the dataset on genes that were annotated across the entire mitogenome were partitioned (Supplementary material: T4). To determine the most appropriate DNA substitution model for each gene, the Akaike Information Criterion test implemented in jModelTest2 (Darriba et al., 2012; Guindon & Gascuel, 2003) was used (see selected models for each gene in Supplementary material: T4). The partitioned mitogenome dataset was used to estimate a Bayesian phylogenetic tree in MrBayes (Ronquist et al., 2012). The Bayesian analysis was carried out by running 4 MCMC chains for 5 million generations and saving the trees every 1,000 generations. The consensus phylogenetic tree was constructed using a 10% burn-in.
Genetic diversity. Individual DNA segments from the coxI, Cytb, ND2, and -ND3 genes were used to estimate the genetic diversity parameters for the A. chrysaetos species complex since these were the genes that had a relative amount of A. chrysaetos DNA sequences available in GenBank. Using DNA sequences from the Baja California individuals produced in this study and reference sequences from the species complex (Supplementary material: T3) the following indices were estimated with DnaSP 6 (Rozas et al., 2017): the number of segregating sites (S), number of haplotypes (Nh), haplotype diversity (Hd), and nucleotide diversity (π). Genetic diversity indices were calculated among all sequences included in the respective datasets and within specific groups of sequences according to phylogenetic clades or geographic origin of samples. In addition, the genetic distance (uncorrected p-distance) between and within the phylogenetic clades observed within the mitochondrial genetic lineages in PAUP software was estimated (Swofford, 1993). The clade OTU1 included sequences from a diverse geographical background, including all the DNA sequences of the Baja California golden eagles. In contrast, the A. chrysaetos DNA sequences grouped in the other closely related clade were classified as OTU2. Additionally, the DNA sequence of an individual of A. chrysaetos from California (Southern Sierra Nevada) was analyzed separately to compare its genetic distance with those from Baja California, given the geographical proximity between both populations.
Results
Raw sequencing reads ranged from 3,324,579 (individual Ach 1) to 539,825 (individual Ach 4). The number of total sequences after the filter ranged between 3,251,746 (Ach 1) and 526,875 (Ach4) (Supplementary material: T2A).
Assembly and annotation. There were large differences in assembly between individuals; for example, the average depth in the reference mapping for individual Ach 1 is 30.5X, while for Ach 6 is 168.3X, even though Ach 6 had fewer total reads than Ach 1. Assemblies for individuals Ach 1-5 showed fewer aligned sequences (Supplementary material: T2A). For de novo assembly, the average depth for individual Ach 6 is 8.0X. The longest scaffold generated (17,472 bp) allowed us to assemble the entire mitogenome of the Ach 6 individual, eliminating the need to align multiple scaffolds to recover the full length (Supplementary material: T2B). The data yield was inferior when doing de novo assembly for Ach 1-5 individuals, and none of these individuals allowed the recovery of the complete mitochondrial genome (Supplementary material: T3B).
Figure 2. Mitogenome annotated by de novo assembly for Golden Eagle (Ach6) from Baja California, Mexico. In the central part of the diagram in a ring form is shown a scale of the length of mitochondrial genome clockwise. Enlarged view of the section showing the division into 2 fractions of the mitochondrial gene for NADH dehydrogenase subunit 3 (ND3). The stop codon of the ND3a fraction is highlighted in red, the insertion of nucleotide A in yellow, and the start codon of the ND3b fraction in green. The ND3 gene is composed entirely of 352 bp; the ND3a fraction is 207 bp long, while the ND3b fraction is 144 bp long.
Based on the above results, we annotated and curated the mitochondrial genome using only the complete mitogenome obtained by de novo assembly of the Ach 6 individual. All 3 annotation tools identified tRNAs, and none of the tools identified introns within the sequence. The MITOS program (Fig. 2, Supplementary material: T4) was the only software that identified all the genetic elements, such as the 22 tRNAs, 2 rRNAs, and the 13 protein-coding genes that make up the mitogenome. This last annotation was the most complete, so the curation was made from it. We found a region with 7 nucleotide bases that is composed of an AGA stop codon of the ND3 protein, followed by an adenine (A) and finally an ATC start codon that codes for isoleucine. This region divides the ND3 fragment into the 2 regions proposed here (ND3a and ND3b, Fig. 2).
Synteny and search for polymorphisms. The linear order of the mitogenome of Ach 6 coincides with the reference genome (Accession number MT319112.1, Supplementary material: T5). We consider the SNPs and INDELs identified in the Ach 6 reliable, given that the average coverage is 168.3X, and their call quality was high (Supplementary material: T2A). Most SNPs and INDELs are found within genes that code for some protein, transfer RNA and ribosomal genes. The coxI gene presented the greatest number of SNPs with respect to the reference genome in the 6 individuals, while the Cytb and tRNA F genes presented the greatest length changes in the INDELs. In addition, in the ND3 gene, there is a consistent change in the sequences of the 6 individuals with respect to the reference genome (Table 1).
Figure 3. Bayesian phylogenetic tree based on the mitogenome for Golden Eagle voucher specimens from northwestern Baja California, Mexico. Posterior probabilities are shown above internal nodes. GenBank accession numbers are positioned next to reference sequences. Aquila nipalensis was used as an outgroup for rooting the tree. Two Operational Taxonomic Units are identified (OTU 1, 2).
Phylogenetic relationships. With the use of the mitogenome we constructed a phylogenetic tree that identified 2 clades with strong support (1): OTU1 that includes all the individuals analyzed in this study, plus 2 GenBank sequences (Aquila chryseatos, LR82062.1 from the United Kingdom and A. heliatica, NC_035806.1), and OTU2 that corresponds to a GenBank sequence of a specimen collected in California (Fig. 3). For their part, the phylogenetic trees obtained with the fragments of individual genes (ND2, ND3, Cytb and coxI) did not resolve the divergence between individuals from Baja California and California obtained with the mitogenome, but all except the ND3 gene showed a divergence (OTUs 1 and 2), mainly with respect to eagles from the European continent (when information on the collection location is reported, Supplementary material: F1).
Genetic diversity. We observed low genetic diversity, for instance, ND3, ND2, and coxI each had a single segregating site (Table 2), resulting in low nucleotide diversity values (0.0005, 0.0008, and 0.0002, respectively). In contrast, Cytb was the most polymorphic, with a total of 10 segregating sites and a nucleotide diversity an order of magnitude larger (0.00297) (Table 2). The genetic diversity found within OTU1 was higher than that found in OTU2 (Table 2), although OTU1 was not a geographically homogeneous clade among the trees of the different mitochondrial genes, for example in the case of ND2 it was made up of North American eagles, including the individuals of this study, but coxI apart from the previous ones was constituted with eagles from Sweden, Norway and Japan (Table 2).
Discussion
Despite using 2 sequencing platforms for the mitogenome of Golden Eagle from northwestern Baja California, Mexico, we successfully assembled the mitochondrial genome of all 6 individuals by the reference method. This included some individuals (Ach 1-5) with non-consensus regions. Bolger et al. (2014) recommend a pre-processing step of the readings before any analysis, be it assembly by reference or de novo, and mention that if the library and identification adapters are not removed, they can be incorporated in the final assembly. In our case, the pre-processing of the sequences positively influenced the performance of both assemblies and the SNP’s call since the quality of the sequence of the 6 individuals was considerably improved after pre-processing in all aspects of quality reporting.
Table 1
Location of SNP’s and INDEL’s in mitogenomes. Call of SNP’s and INDEL’s for each the Ach 1-6 specimens versus the reference genome of Golden Eagle from the southern Sierra Nevada in California, USA (GenBank accession: KF905228). Pos: Mitogenome position; Ref: nucleotide present in reference sequence. ψCR: Mitochondrial pseudogene control region.
| Pos. | Ref. | Ach1 | Ach2 | Ach3 | Ach4 | Ach5 | Ach6 | Gen |
| SNP’s | ||||||||
| 1,505 | G | A | A | A | ||||
| D-loop | ||||||||
| 1,515 | T | C | ||||||
| 3,130 | G | A | A | A | ||||
| 3,566 | A | G | ND6 | |||||
| 3,695 | T | A | A | A | ||||
| 3,696 | T | A | A | A | ||||
| 5,557 | G | A | A | A | A | A | LSU rRNA | |
| 6,646 | G | A | A | A | ND1 | |||
| 7,141 | A | G | ||||||
| 8,269 | C | T | A | T | ND2 | |||
| 8,593 | A | G | G | A | G | |||
| 8,723 | G | A | A | A | A | tRNA-Trp | ||
| 10,925 | C | T | T | T | T | T | ||
| 10,963 | T | C | ||||||
| 10,970 | C | T | T | T | T | T | ||
| 10,975 | T | C | coxII | |||||
| 11,003 | A | G | G | |||||
| 11,011 | C | T | ||||||
| 11,205 | A | G | ||||||
| 11,213 | C | T | ||||||
| 11,237 | C | T | ||||||
| 11,282 | G | A | ||||||
| 11,284 | G | T | ||||||
| 11,285 | A | C | ||||||
| 11,351 | T | C | ||||||
| 11,354 | A | G | ||||||
| 11,378 | C | T | T | |||||
| 11,432 | C | T | T | T | T | T | ||
| 11,489 | A | C | C | C | C | C | C | |
| 11,493 | T | T | tRNA-Lys | |||||
| 11,496 | G | A | ||||||
| 11,505 | T | G | ||||||
| 17,278 | G | A | A | ND5 | ||||
| INDEL’s | ||||||||
| Table 1. Continued | ||||||||
| Pos. | Ref. | Ach1 | Ach2 | Ach3 | Ach4 | Ach5 | Ach6 | Gen |
| 3 | CTAA | CTAA CTTC CAAA CTAA | Cyt b | |||||
3,550 | T | TGTG AACA A | ψCR | |||||
| 3,701 | CAAA | CCCA CCAA TA | CCAA CAAT AT | tRNA-Phe | ||||
| 13,378 | T | TC | TC | TC | TC | TC | TC | ND3 |
Table 2
Genetic diversity and haplotype composition from gene segments. Tax Set: Group of sequences included in taxonomic set; N: number of sequences; bp: base pairs of DNA sequence included in alignment; S: number of segregating sites; Nh: number of haplotypes; Hd: haplotype diversity; Nd: nucleotide diversity; North America: sequences from either Canada or USA. All sequences from ND3 formed a single clade, and thus were analyzed as a single group. OTUs 1 and 2 in each individual gene refer to the clades detected in each phylogenetic analysis, see Supplementary material F1.
| Mitochondrial sequence | Tax Set | N | bp | S | Nh | Hd | Nd |
| ND3 | All | 11 | 352 | 1 | 2 | 0.182 | 0.00053 |
ND2 | OTU1 | 8 | 1,039 | 0 | 1 | 0 | 0 |
| OTU2 | 4 | 1,039 | 0 | 1 | 0 | 0 | |
| North America | 2 | 1,039 | 0 | 1 | 0 | 0 | |
| Europe | 3 | 1,039 | 0 | 1 | 0 | 0 | |
| All | 12 | 1,039 | 1 | 2 | 0.485 | 0.00087 | |
coxI | OTU1 | 13 | 1,551 | 1 | 2 | 0.154 | 0.00025 |
| OTU2 | 2 | 1,551 | 0 | 1 | 0 | 0 | |
| North America | 4 | 1,551 | 0 | 1 | 0 | 0 | |
| Europe | 4 | 1,551 | 1 | 2 | 0.5 | 0.00082 | |
| All | 15 | 1,551 | 1 | 2 | 0.133 | 0.00022 | |
Cytb | OTU1 | 13 | 1,143 | 1 | 2 | 0.154 | 0.00016 |
| OTU2 | 3 | 1,143 | 1 | 2 | 0.667 | 0.00069 | |
| North America | 4 | 1,143 | 1 | 2 | 0.5 | 0.00052 | |
| Europe | 5 | 1,143 | 9 | 3 | 0.02688 | 0.00541 | |
| All | 16 | 1,143 | 10 | 4 | 0.442 | 0.00297 |
The most reliable mitogenome assembly was that of Ach 6 individual due to its higher alignment rate, its greater depth, and its 100% coverage of the reference genome. This ensemble can be considered true and not an artifact due to the procedures used, such as sampling every 500 generations and a burn-in value of 600,000 generations (Lerner & Mindell, 2005). In addition, we analyzed the data with Bowtie2, which implements an alignment strategy based on the FM-Index and Burrows-Wheeler. These programs have been shown to work well for applications such as INDEL discovery (Lindner & Friedel, 2012) and for aligning long sections of the reference genome (Thankaswamy-Kosalai et al., 2017). Hunt et al. (2014) pointed out that the best result of assembling a genome de novo is when it is contained in a single scaffold for the entire mitochondrial DNA molecule or for each chromosome in the case of the nuclear genome. The Golden Eagle mitochondrial genome assembly of individual Ach 6 was the best because it was found within a single scaffold (17,472 bp) and its depth was at least 8.0X (Supplementary material: T2B; Baker, 2012).
The extension and position of the genes in the mitogenome of the Golden Eagle described in this study (Fig. 2, Supplementary material: T4) showed differences compared to those reported by Doyle et al. (2014) from an individual sampled in central California. For example, the size of the mitogenome described here is larger (17,472 bp) than the one reported by Doyle et al. (2014) (17,332 bp), with a difference of 140 bp. When experimentally analyzing the size of the ND3 gene amplified fragments, the size was between 700 and 800 bp, that is, the theoretically expected size. Another difference consisted in the fraction of the gene encoding the ND3 protein reported here as split into 2 segments (ND3a and ND3b). The distinction of the 2 subunits is, within the reading frame, a region made up of 7 nucleotides between the ATC start codon that starts the protein, and the TAA stop codon that marks the end of the protein. The first 3 of these 7, code for a new AGA stop codon, followed by an adenine (A) nucleotide and 3 nucleotides that code for a new ATC start codon. This causes 2 coding portions to coexist, annotated as ND3, separated by a single nucleotide that changes the reading frame, causing the second portion to appear with its start and stop codons. Mindell et al. (1998) report a nucleotide that is not translated within the ND3 protein sequence. Eberhard and Wright (2016) mentioned that such a trait is observed in the entire order Psittaciformes (parrots, parakeets, and allies).
Specifically, a variety of distinct indels have been found within several of the mitochondrial protein-coding genes in Psittaciformes (Eberhard & Wright, 2016), with variations in terms of their evolutionary origin since some are present in all Psittaciformes and some are more recent in origin and only found within specific taxa. Slack et al. (2003) reported variation in the length of the mitochondrial ND6 gene in other avian taxa. Overall, the biological implications of these variations remain unclear until more studies on the proteins are performed to determine if the protein compositions become altered by these variations. However, the experimental verification carried out in the present study strengthens our annotation as 2 fractions for the ND3 region of the A. chrysaetos mitogenome. Also, we noted the regulatory non-coding region (the pseudo control region, ψCR) to be highly conserved among the A. chrysaetos individuals analyzed, which is a relatively conserved region within the order Accipitriformes (cf. Song et al., 2015) and is also found in the same position between the tRNA -E and tRNA-F, as shown by Liu et al. (2017) for the mitogenome of Accipiter gularis, another congeneric species.
Synteny and search for polymorphisms. The assembled Golden Eagle mitogenome did not present changes in gene order with respect to the reference (KF905228.1, a male Golden Eagle from southern Sierra Nevada, California). As expected, neither did it present changes in the order of the genes with respect to other species of the same group of birds (Accipitridae), such as Buteo buteo (Haring et al., 2001), Accipiter virgatus (Song et al., 2015), Aquila fasciata (Jiang et al., 2015) and Accipiter gularis (Liu et al., 2017). In addition, it is consistent with one of the general models of the birds’ mitogenome that is characterized by a duplication of the control region and other adjacent genes, which were subsequently degraded, giving rise to the ψCR (Eberhard & Wright, 2016). SNP’s and INDEL’s observed in the mitogenome of the Ach 6 individual meet the needed quality (probability of not being an error) and sequence depth (Li, 2011), so they can be considered as potential markers that should be tested in future population studies, considering some technical aspects, and thus avoid the call of false positives (The 1000 Genomes Project Consortium, 2015).
Reconstruction of phylogenetic relationships. The phylogenetic tree classified all Golden Eagle individuals into 2 mitogenome clades (OTU1 and OTU2). All individuals from Baja California, an individual from the United Kingdom, and an A. heliaca individual clustered within OTU1 (Fig. 3), while OTU2 was composed of a single Golden Eagle individual from California. The genetic distance between the 2 OTUs was 0.002 nucleotide differences per site (Supplementary material: T6). This pattern is not surprising given the poor sampling of taxa for both the mitogenomes and individual genes of A. chrysaetos. Once additional mitogenomes are available for this species, it will become possible to increase the phylogenetic resolution that could confirm or reject the potential presence of multiple mitochondrial lineages circulating within Golden Eagle populations. Nevertheless, given the broad geographic area included in the mitogenome dataset (i.e., an A. chrysaetos individual from the United Kingdom, as well an A. heliaca individual), and the low genetic diversity observed within these mitochondrial sequences (see below), suggests that at the mitochondrial level, the genetic diversity found within A. chrysaetos populations is low.
The use of genes or fragments of the mitochondrial genome allows for the analysis to include a larger data set in the number of locations and individuals. The topology of 3 of the 4 individual gene trees (ND2, Cytb and coxI) did support the presence of at least 2 mitochondrial clades (Supplementary material: F1). However, these gene trees were not consistent when grouping the North American Golden Eagle individuals on the same OTUs.
Future studies targeting molecular markers with a higher mutational rate (i.e., microsatellites) or genomics approach could potentially uncover the presence of genetic differences that are not well defined at the mitochondrial level but might be biologically important. One such difference is seen in the species’ bivalent behavior and its migration pattern in the west (Mcintyre & Lewis, 2016) and northwest of the USA, and through Canada (Bedrosian et al., 2018). The mitogenomic lineages that potentially infer the difference between the eagles of Baja California and central California could correspond to the area where 2 ecological populations are coexisting (De León-Girón et al., 2016). The first population is migratory, originating in western North America (Oregon), while the second lineage is resident in Baja California. According to Craig et al. (2016), this lineage has 1 of the less frequent haplotypes of the mitochondrial control region and is restricted to California. The ability of the Golden Eagle to adapt to different environments (Judkins & Van Den Bussche, 2018), the large dispersal distances of reproductive or floating between Mexico and the USA (De León-Girón et al., 2016, 2024; Rodríguez-Estrella et al., 2020; Tracey et al., 2017), and its extensive home range of reproductive pairs (D’Addario et al., 2019), would support these hypotheses. Besides, the presence of these groups of individuals (reproductive and non-reproductive) in Baja California would be part of the reproductive behavior of the species (Watson et al., 2011), that is, the process of succession and substitution of reproductive pairs in the region (De León-Girón et al., 2016).
Genetic diversity. The values of mitochondrial genetic diversity calculated for most genes (except Cytb) were near zero (Table 2). We expected this result, considering that they are genes that code for proteins with highly conserved regions (Dawnay et al., 2007) and are subject to biochemical limitations that cause high levels of homoplasy (Faria et al., 2007). However, Bates et al. (2003) mentioned that ND2 is genetically more diverse than Cytb, while Faria et al. (2007) stated that ND2 is one of the most variable mitochondrial genes within birds and, therefore, regularly implemented in population genetics. A pattern we did not observe in A. chrysaetos since Cytb was the gene with the highest genetic diversity observed, with a haplotype diversity of 0.154 and 0.667 for OTU1 and OTU2, respectively, while both OTUs had a haplotype diversity of zero for the ND2 gene (Table 2).
Implications for conservation. Golden eagles are recognized for presenting large expanses of territory, and their resident, migratory and floating populations (reproductive adults without territories), promote the population gene flow (Craig et al., 2016; Poessel et al., 2022). The mitochondrial DNA data produced in this study confirm a very close genetic relationship between the northwestern Baja California individuals and those from the USA. Therefore, the execution of bi-national conservation programs by the agencies of both countries (SEMARNAT and USFWS) are priorities for the “Californian-Baja Californian” metapopulation of Golden Eagle. It is necessary to increase the number of young individuals with satellite tracking in the southern Baja California peninsula, to continue the genetic characterization (with different types of markers, mitochondrial DNA, microsatellites and SNPs obtained by Next Generation Sequencing) and population monitoring to evaluate the structure and connectivity of the species in both countries.
Our findings warrant developing joint conservation efforts between the governments of Mexico and the USA to monitor and preserve the North American Golden Eagle.
Acknowledgements
Funding for the sequencing was covered by the SAGARPA-INAPESCA PIDETEC project 2017/0647. The authors thank Travis Glen and Natalia Juliana Bayona Vásquez for their Illumina sequencing services. We thank 3 anonymous reviewers who helped improve the manuscript.
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Avifauna of cloud forest riparian corridors in a degraded landscape in eastern Mexico
Omar A. Hernández-Dávila a, *, Vinicio Sosa b, Javier Laborde b
a Instituto de Biotecnología y Ecología Aplicada, Av. de las Culturas Veracruzanas No. 101, Col. Emiliano Zapata, 91090 Xalapa, Veracruz, Mexico
b Instituto de Ecología, A.C., Red de Ecología Funcional, Carretera Antigua a Coatepec, No. 351, El Haya, 91073 Xalapa, Veracruz, Mexico
*Corresponding author: omar.hernandez.davila@outlook.com (O.A. Hernández-Dávila)
Abstract
Cloud forests are known for their remarkable biodiversity and provide many ecosystem services. However, this biodiversity is in jeopardy due to the conversion of forests to other land uses. At its northernmost range in the Neotropics, cloud forest persists in remnant fragments immersed in an agricultural matrix that still has arboreal elements, such as riparian corridors. In this study we characterize the avifauna present in cloud forest riparian corridors in a highly degraded landscape of Mexico. We classified the avifauna in terms of migratory and conservation status, trophic guild, body mass, forest stratum and habitat preference. In 14 riparian corridors we recorded 86 bird species (75% were resident). Insectivorous and frugivorous species represented 79% of total richness. Almost 65% of species prefer the mid-story or canopy forest strata, while 46% were habitat generalists. Despite crossing open agricultural areas, cloud forest riparian corridors still harbor a diverse assemblage of bird species, which includes not only those tolerant to disturbance, but also species that are typical of old-growth cloud forest. We suggest that these remnants may be crucial for forest birds that move across the fragmented landscape.
Keywords: Riparian strips; Bird community; Fragmented landscape; Reservoirs
Avifauna de corredores ribereños de bosque de niebla en un paisaje degradado del este de México
Resumen
Los bosques de niebla son conocidos por su notable biodiversidad y sus servicios ecosistémicos. Sin embargo, esta biodiversidad está en peligro debido a su conversión a otros usos de suelo. En su distribución más septentrional en el Neotrópico, el bosque de niebla persiste en remanentes inmersos en una matriz agrícola que aún conserva elementos arbóreos, como los corredores riparios. En este estudio caracterizamos la avifauna presente en dichos corredores en un paisaje altamente degradado de México. Clasificamos a la avifauna en términos de su estatus migratorio y de conservación, gremio trófico, masa corporal y preferencia de hábitat y estrato forestal. En 14 corredores riparios registramos 86 especies de aves. Las especies insectívoras y frugívoras representaron 79% de la riqueza total. Casi 65% de las especies prefieren los estratos forestales medios o de dosel, mientras que 46% fueron generalistas de hábitat. Aunque los corredores riparios atraviesan zonas agrícolas, siguen albergando un conjunto diverso de aves, que incluye no solo aquellas tolerantes a las perturbaciones, sino también especies características del bosque de niebla conservado. Sugerimos que estos remanentes pueden ser cruciales para las aves forestales que se desplazan por el paisaje fragmentado.
Palabras clave: Franjas riparias; Comunidad de aves; Paisaje fragmentado; Reservorios
Introduction
Tropical montane cloud forest (hereafter, cloud forest) is one of the most important terrestrial ecosystems worldwide. It provides environmental services such as carbon sequestration and water capture, and also mitigates flooding and drought (Bruijnzeel et al., 2011). Cloud forest is among the most biodiverse ecosystems in the world and hosts a remarkable diversity of flora and fauna in which spatial variation is prominent (i.e., high beta diversity), as well as a high proportion of endemic species (Aldrich et al., 2000; Karger et al., 2021). Mexican cloud forests are particularly rich in species of trees, shrubs and epiphytes, along with amphibians, reptiles, birds and mammals (Gual-Díaz & Rendón-Correa, 2014). It is estimated that Mexican cloud forests are home to 551 bird species (i.e., 50% of the total richness of the avifauna in the country), and the cloud forests in the state of Veracruz are home to 346 bird species (Navarro-Sigüenza et al., 2014). The richness of cloud forest avifauna in other parts of the country ranges from 196 to 335 species per region (Hernández-Baños et al., 1995; Navarro-Sigüenza et al., 2014). Despite its relevance, Mexican cloud forest is currently in jeopardy, not only because of agricultural expansion and uncontrolled urban growth, but also due to the illegal extraction of its species and unsustainable use of its resources (Toledo-Aceves et al., 2011). In the central part of Veracruz in the 1990s, it was estimated that there were 426 km² of cloud forest, an area that was reduced to 279 km² by 2003 (Muñoz-Villers & López-Blanco, 2008). Even though the rate of Mexican cloud forest deforestation has slowed over the last decade, today it is estimated to cover only 175 km² in Veracruz (Bonilla-Moheno & Aide, 2020). Cloud forest remnants are currently found as numerous fragments of different sizes (1 to 30 ha, usually), which are immersed in an extensive agricultural matrix (Williams-Linera et al., 2002) that still contains distinct arboreal elements such as isolated trees, living fences and forested riparian corridors. The presence of these arboreal elements within the agricultural matrix, which also include small patches (< 5 ha) of secondary forest, could be relevant to the maintenance of different groups of native flora and fauna in anthropic landscapes (Toledo-Aceves et al., 2014).
Cloud forest riparian corridors are single rows of trees growing along each side of permanent streams or rivers that have been left uncut by farmers when converting the forest to crop fields or pastures. These narrow, elongated belts or strips of tall trees have a dense woody undergrowth (riparian corridors, hereafter) and extend along rivers for kilometers. They are usually the most conspicuous arboreal element in agricultural landscapes. Forested riparian corridors that cross agricultural plots provide several environmental services, such as riverbank stabilization, nutrient recycling and enrichment, and filtering and retention of agrochemical pollutants in surface runoff, among other services (Cole et al., 2020). Additionally, riparian corridors contribute to the conservation of several taxonomic groups in anthropic landscapes, including native woody plant species (Hernández-Dávila et al., 2020), amphibians (Rodríguez-Mendoza & Pineda, 2010), bats and non-flying mammals (Griscom et al., 2007; Zarazúa-Carbajal et al., 2017). For birds, forested riparian corridors have been shown to function as refuges, foraging and nesting sites, as well as habitat corridors or stepping stones for moving across fragmented landscapes (Domínguez-López & Ortega-Álvarez, 2014; Kontsiotis et al., 2019; Lees & Peres, 2008). In lowland tropical regions that have been converted to agriculture, relatively wide riparian corridors connected to large forest remnants have been found to harbor a higher richness and abundance of forest birds than narrower and unconnected riparian corridors (Arizmendi et al., 2008; Domínguez-López & Ortega-Álvarez, 2014; Pliscoff et al., 2020). The presence of linear arboreal elements within the agricultural matrix might help to maintain bird diversity in anthropic landscapes (de Zwaan et al., 2022). To date, the majority of studies on the avifauna that uses riparian corridors in anthropic or/and fragmented landscapes has been done in lowland areas originally covered by tropical rainforest or seasonally dry tropical forest (Graham et al., 2002; Latta et al., 2012; Villaseñor-Gómez, 2008). In particular, for riparian strips of tropical montane cloud forest, Hernández-Dávila et al. (2021) found that landscape composition (i.e., urban and forest area) and vegetation structure (mean height of vegetation) positively influence the richness and abundance of generalist and specialist birds using riparian strips. However, the bird community of cloud forest riparian habitats that could potentially be using these remnants in anthropic landscapes has not been characterized to date. In this sense, taking into account that the original area covered by cloud forest has been drastically reduced by human activities and that the remaining forest fragments are surrounded by an extensive agricultural matrix, in which still there are arboreal riparian strips, we wanted to determine and characterize the composition and structure of the bird community present in these remnants. The latter, in order to find out if these riparian strips can serve as biodiversity reservoirs or landscape connectors, that facilitate the movement of birds across the landscape and which bird species use them. The objectives of this paper were: 1) to determine the richness, diversity and composition of birds that use riparian corridors of cloud forest in a highly modified anthropic landscape, and 2) to characterize the avifauna that uses these corridors in terms of their conservation and migratory status, trophic guild, and habitat preference. To date, no characterization of the bird community present in riparian corridors of cloud forest has been carried out. Given that the cloud forest is fragmented and natural or semi-natural remnants, such as arboreal riparian corridors, can support different groups of flora and fauna, it is necessary to determine and analyze the avifauna present in these corridors. This knowledge is needed to design and implement management strategies of riparian corridors and improve the odds of native cloud forest species conservation in current landscapes.
Materials and methods
The study area is located in the central part of the state of Veracruz, Mexico in the upper basin of the La Antigua River within 19°31’59”-19°22’42” N, 97°05’36”-96°57’43” W (Fig. 1). The original vegetation was cloud forest (i.e., Tropical Montane Cloud Forest) with an average annual temperature of 18 °C and total annual precipitation from 1,500 to 2,000 mm/year. Fourteen riparian corridor sites (with elevations from 1,190 up to 1,780 m asl) were selected for bird sampling. The sites selected are near to the cities of Xalapa, Coatepec, and Xico and are part of a highly fragmented landscape and are representative of the riparian corridors in the region (i.e., narrow linear bands of remnant cloud forest 2 to 5 m wide growing on both sides of permanent rivers). Since riparian corridors can be several kilometers long, we delimited each of our 14 sampling sites as a riparian tract or segment approximately 400 m long (± 16 m, s.e.) with a continuous tree canopy (i.e., uninterrupted arboreal cover). The 14 riparian segments selected were all separated by more than 1 km. Most of the segments of riparian corridors that we selected for bird sampling cross open cattle pastures, with a few of them adjoining small patches (< 4 ha) of secondary old-growth forest or different types of crop fields (mostly maize or shaded coffee plantations). Dominant tree species in the sampling sites include Platanus mexicana, Liquidambar styraciflua, Palicourea padifolia, Styrax glabrescens, and Perrottetia longistylis. The canopy of these riparian corridors is formed by tall trees usually 15 to 20 m in height, with some surpassing 30 m, however, the average tree height in the sampling sites was 6.3 (± 7.0) m. See Hernández-Dávila et al. (2020) for more details on the vegetation structure and composition of the sites sampled.
The richness and number of birds visiting each sampling site were recorded at 2 fixed point counts set along each segment at least 250 m apart (Gregory et al., 2004). Field observations were conducted in October 2017, and in January, April and July 2018, for a total of 4 visits per point over the course of 1 year, with a 75-day interval between visits to each point. For each point count, all bird sightings were recorded with binoculars over the course of 15 minutes and within 35 m distance by one of us between sunrise and 10:30 am, except on rainy days. It took 3 days to complete the 28 points of a given period. The starting point during each period was alternated to cover the whole morning schedule of observation at each point. Only birds that were perching on the woody vegetation or on the ground or riverbank were recorded. Those flying overhead or perching in open areas nearby outside the riparian corridor were not counted. Thus, a total of 112 counts (14 sites × 2 points/site × 4 visits/site) were done at 28 points, totaling 28 h of observation. In addition to the visual records, we also set mist nets to capture birds to record understory birds visiting some of the sampled corridors, but owing to time and monetary constraints, we were only able to place the nets in 6 of the 14 riparian corridor sites. These 6 sites were chosen randomly and in each of them a total of 15 nets (10 × 2.5 m, each) were set parallel to the river flow on both riversides and at least 50 m apart along the sampled segment. From August 2016 to July 2017, one of the 6 riparian corridors was selected each month for mist-netting. Nets were left in place for 3 consecutive days avoiding rainy days and opened twice a day from sunrise to 11:00 and from 16:00 to sunset. This was done at each site twice over the sampled year. Sampling effort was 1,743 h and 2,250 m² of nets (25 m²/net × 15 nets × 6 sites) in 2 sampling periods at each site. Nets were inspected by 2 people every 30 minutes or less, depending on capture intensity. Each captured bird was marked by trimming a notch at the tip of one of the tail feathers to recognize recaptured individuals. Birds were released in situ immediately after sexing, weighing, and measuring (tarsus, tail, and total length; wing chord length; beak length, width, and depth).
Figure 1. A, Location of the 14 segments of cloud forest riparian corridors (black circles) selected for bird sampling in central Veracruz, Mexico. The main river or streams (lines) and cities (polygons) are shown in gray. Sampling sites: Agua Bendita (AB), Agüita Fría (AF), Acuario (AC), Trucha Feliz (TF), Mariano Escobedo (ME), Truchas Martin (TM), Granada (GR), Trianon (TR), Marina (MA), Rio Matlacobatl (RM), Puente de Dios (PD), Monte Grande (MG), Tlalchy (TL) and Vista Hermosa (VH); B, image (Google Earth – Pro V 7.3.6.9345) is a close-up of 4 of the riparian corridor segments sampled (VH, MG, TL, PD); C, close-up of the PD site, with the river indicated by a blue line. Map by O. Hernández-Dávila.
Bird species were identified using the Sibley (2000) and Howell and Webb (1995) field guides. Nomenclature follows the IOC World Bird List checklist (Gill et al., 2024). Recorded birds were classified as either migratory from North America or resident species, and we also noted whether they were endemic, following Navarro-Sigüenza et al. (2007). The conservation status of each species was based on Mexican federal law NOM-059 (Semarnat 2010), and The IUCN Red List (IUCN, 2024). For trophic guild, each species was classified as: carnivore, insectivore, frugivore, granivore, nectarivore and omnivore (González-Salazar et al., 2014). Additionally, species were categorized into 3 size categories (body mass): small (< 40 g), medium (40 – 100 g) and large birds (> 100 g). Forest stratum preferences were based on Martínez-Morales (2001) and classified as: canopy species (those associated with the upper forest stratum, > 10 m above the ground); midstory species (those that preferentially use forest stratum between 5 and 10 m above the ground); midstory and canopy species (those using both midstory and canopy strata); understory species (those that frequent the forest floor up to 5 m above the ground); understory and midstory species (those using both understory and midstory strata); and all strata species (those that use all forest strata). Finally, main habitat preference was also based on Martínez-Morales (2001) and noted as: forest interior species (those which prefer forest sites away from the forest edge); forest edge species (those that preferentially use the fragment edge less than 100 m away from open areas); forest generalist species (those that are common both at the edge and the interior of forest fragments); and vegetation matrix species (those associated with the agricultural matrix or open areas). For species not reported by Martínez-Morales (2001) we did not assign categories for forest stratum preference and habitat preference. For these cases, species were classified as unknown.
The total richness recorded in riparian corridors was determined by the sum of records obtained from point records and those from mist-nets. Each bird species was characterized in relation to its migratory and conservation status, size, trophic guild, and habitat and forest stratum preferences. Due to the differences in the number of riparian corridors sampled by each sampling method (i.e., 14 for count points and 6 for mist nets) and their differences in data obtained from each method, for all the remaining analyses, only data from count points were taken into account: species accumulation curve and sample coverage were estimated to assess the sample completeness. Hill numbers were calculated to analyze the diversity of bird species in terms of effective number of species (q0), effective number of abundant species or Shannon diversity (q1) and effective number of dominant species or Simpson diversity (q2) (Hsieh et al., 2016; Jost, 2006). The rank-abundance curve of overall recorded avifauna was drawn to show graphically the structure of the bird community (Kindt & Coe, 2005). Analyses of accumulation curves, Hill numbers and range-abundance curves were performed for all riparian strips combined and separately. Finally, similarity in species composition among the 14 sampling sites was estimated using the Jaccard index distance, which varies from 0 to 1. Zero indicates that no species are shared between compared sites, and 1 indicates that species composition is identical between the sites. All analyses were run in R software using the vegan (Oksanen et al., 2020) and BiodiversityR (Kindt, 2021) packages.
Results
We recorded a total of 86 bird species in the 14 riparian corridors. These 86 species belonged to 9 orders, 27 families and 67 genera (Table 1). The richest families were Parulidae (16 species); Tyrannidae (13 spp.); Trochilidae (9 spp.) and Turdidae (6 spp.). Of the 86 bird species, there were 67 recorded at the point-counts and 48 trapped in the nets, with 29 recorded with both field methods (Table 1).
Of the 86 recorded species, 65 were resident species and 21 were migratory. Only 2 of the recorded species are endemic to Mexico: Melanotis caerulescens (Blue Mockingbird) and Cardellina rubra (Red Warbler). Four species are under special protection status by Mexican federal law (Semarnat, 2010): Accipiter striatus (Sharp-shinned Hawk), Cinclus mexicanus (American Dipper), Psarocolius montezuma (Montezuma Oropendola) and Catharus mexicanus (Black-headed Nightingale-thrush), and 1 species is threatened: Catharus frantzii (Ruddy-capped Nightingale-thrush). One species, Selasphorus rufus (Rufous Hummingbird), is regarded as Near Threatened in the IUCN list. The richest trophic guilds were insectivorous (63% of species) and frugivorous birds (15%) (Fig. 2A). Regarding forest stratum preference, 18% of the species recorded prefer the midstory to canopy strata, and 17% are midstory specialists (Fig. 2B). For habitat preference, habitat generalists were the most strongly represented accounting for 30% of richness, followed by forest-interior species with 18% (Fig. 2C). Most recorded species (66%) were relatively small with a body mass < 40 g, and only 12% of species were heavier than 100 g (Fig. 2D).
During the 28 point counts we recorded a total of 816 detections of 67 bird species using the corridors. In the mist nets placed in 6 of the corridors, we captured 273 birds belonging to 48 species. The species accumulation curve for all corridors reached a sample completeness of 97%. According to Hill numbers the bird diversity was of 67 species observed (q0), 18 common species (q1) and 8 dominant species (q2) (Fig. 3). Individually, the sample completeness for each of the 14 riparian segments varied between 79% and 93% per segment; and regarding diversity, q0 ranged from 9 to 24 bird species per segment, while q1 ranged from 5 to 13 species and q2 from 2 to 13 species (Fig. 4).
Table 1
List of the 86 bird species recorded in 14 cloud forest riparian corridors in the anthropic landscape of central Veracruz, Mexico. Shown are the number of visual detections (# detections) of each species estimated by point counts and the number of birds
captured in mist nets. Seed dispersing birds are indicated with an asterisk (*). Species are ordered by rank (only for data obtained by point counts). We also show for each species: migratory status, resident (R), North American migrant (M). Protection status: threatened (A), special protection (Pr), least concern (LC), near threatened (NT). Trophic guild: insectivore (In), frugivore (Fr), nectarivore (Ne), omnivore (Om), granivore (Gr), carnivore (Ca). Size (body mass in g): small < 40 g (1), medium-sized 40 – 100 g (2), large > 100 g (3). Habitat preference: forest interior (FI), forest edge (FE), forest generalist (FG), vegetation matrix (VM), and no information available (-). Forest stratum preference: understory (U), understory and midstory (UM), midstory (M), midstory and canopy (MC), canopy (C), and no information available (-). See below for bibliographic sources.
| Species | # detections by point counts | # captures in mist nets | Migratory status | Protection status | Trophic guild | Size category | Habitat preference | Stratum preference |
| Chlorospingus flavopectus* | 234 | 17 | R | -/LC | In-Fr | 1 | FG | MC |
| Cardellina pusilla | 102 | 15 | M | -/LC | In | 1 | FG | A |
| Psilorhinus morio* | 79 | 2 | R | -/LC | Om | 3 | FG | MC |
| Myadestes occidentalis* | 51 | 10 | R | -/LC | Fr | 2 | FG | MC |
| Empidonax difficilis | 36 | 18 | R | -/LC | In | 1 | FG | M |
| Myiozetetes similis* | 33 | 3 | R | -/LC | In-Fr | 1 | VM | C |
| Parkesia motacilla* | 20 | 14 | M | -/LC | In-Fr | 1 | FI | U |
| Psarocolius montezuma* | 20 | – | R | Pr/LC | Fr | 3 | – | – |
| Setophaga townsendi | 12 | – | M | -/LC | In | 1 | FG | MC |
| Sayornis nigricans | 11 | 2 | R | -/LC | In | 1 | – | – |
| Tityra semifasciata* | 9 | – | R | -/LC | In-Fr | 2 | – | – |
| Turdus assimilis* | 9 | 6 | R | -/LC | Fr | 2 | FE | M |
| Turdus grayi* | 9 | 6 | R | -/LC | Fr | 2 | FE | M |
| Cinclus mexicanus | 8 | – | R | Pr/LC | Ca | 2 | – | – |
| Leiothlypis ruficapilla | 8 | 1 | M | -/LC | In | 1 | FI | MC |
| Mitrephanes phaeocercus | 8 | – | R | -/LC | In | 1 | FG | M |
| Quiscalus mexicanus* | 8 | 1 | R | -/LC | Om | 3 | – | A |
| Vireo solitarius* | 8 | 3 | M | -/LC | In-Fr | 1 | FG | M |
| Henicorhina leucophrys* | 7 | 4 | R | -/LC | In-Fr | 1 | FG | U |
| Melanerpes aurifrons* | 6 | – | R | -/LC | In-Fr | 2 | – | – |
| Thraupis abbas* | 6 | 1 | R | -/LC | Fr | 2 | FE | C |
| Contopus pertinax | 5 | 1 | R | -/LC | In | 1 | FG | C |
| Cyanolyca cucullata* | 5 | – | R | -/LC | Om | 3 | FI | MC |
| Euphonia hirundinacea* | 5 | 9 | R | -/LC | Fr | 1 | – | – |
| Lepidocolaptes affinis | 5 | 3 | R | -/LC | In | 1 | FG | M |
| Mniotilta varia | 5 | 1 | M | -/LC | In | 1 | FI | MC |
| Myioborus miniatus | 5 | 1 | R | -/LC | In | 1 | FI | M |
| Chloroceryle americana | 4 | – | R | -/LC | Ca | 2 | FE | MC |
| Empidonax hammondii | 4 | – | M | -/LC | In | 1 | – | – |
| Melanerpes formicivorus* | 4 | – | R | -/LC | In-Fr | 2 | FE | MC |
| Piranga leucoptera* | 4 | – | R | -/LC | In-Fr | 1 | – | – |
| Table 1. Continued | ||||||||
| Species | # detections by point counts | # captures in mist nets | Migratory status | Protection status | Trophic guild | Size category | Habitat preference | Stratum preference |
| Catharus frantzii* | 3 | 4 | R | A/LC | Fr | 1 | FI | UM |
| Momotus coeruliceps* | 3 | 2 | R | -/LC | In-Fr | 3 | FG | M |
| Myiarchus tuberculifer* | 3 | 1 | R | -/LC | In-Fr | 1 | FG | M |
| Polioptila caerulea | 3 | – | M | -/LC | In | 1 | FG | MC |
| Ptiliogonys cinereus* | 3 | – | R | -/LC | In-Fr | 1 | FG | C |
| Rupornis magnirostris | 3 | 1 | R | -/LC | Ca | 3 | – | – |
| Vireo cassini* | 3 | – | M | -/LC | In-Fr | 1 | – | – |
| Amazona albifrons | 2 | – | R | -/LC | Gr | 3 | – | – |
| Basileuterus belli* | 2 | 11 | R | -/LC | In-Fr | 1 | FI | UM |
| Basileuterus rufifrons | 2 | – | R | -/LC | In | 1 | FG | UM |
| Chlorophonia elegantissima * | 2 | – | R | -/LC | Fr | 1 | FI | MC |
| Dendrocincla homochroa | 2 | 1 | R | -/LC | In | 2 | – | – |
| Leptotila verreauxi | 2 | – | R | -/LC | Gr | 3 | FG | U |
| Pyrocephalus rubinus | 2 | – | R | -/LC | In | 1 | – | – |
| Saltator atriceps | 2 | 2 | R | -/LC | Fr | 2 | – | – |
| Setophaga tigrina | 2 | – | M | -/LC | In | 1 | – | – |
| Setophaga virens | 2 | – | M | -/LC | In | 1 | FI | MC |
| Sporophila morelleti | 2 | 1 | R | -/LC | Gr | 1 | – | – |
| Campylorhynchus zonatus | 1 | – | R | -/LC | In | 2 | – | – |
| Cardellina rubra | 1 | – | R | -/LC | In | 1 | – | – |
| Catharus mexicanus* | 1 | 12 | R | Pr/LC | Fr | 1 | FG | U |
| Corthylio calendula | 1 | – | M | -/LC | In | 1 | FG | U |
| Dives dives* | 1 | 2 | R | -/LC | In-Fr | 2 | VM | A |
| Dumetella carolinensis* | 1 | 2 | M | -/LC | In-Fr | 2 | VM | M |
| Empidonax minimus | 1 | – | M | -/LC | In | 1 | – | – |
| Icterus bullockii* | 1 | – | R | -/LC | In-Fr | 1 | FE | M |
| Megarynchus pitangua* | 1 | – | R | -/LC | In-Fr | 2 | – | – |
| Ortalis vetula* | 1 | – | R | -/LC | Fr | 3 | FE | M |
| Pachyramphus aglaiae | 1 | – | R | -/LC | In | 1 | FI | MC |
| Piaya cayana* | 1 | – | R | -/LC | In-Fr | 3 | FG | M |
| Piranga flava* | 1 | – | R | -/LC | In-Fr | 1 | FE | MC |
| Piranga rubra* | 1 | – | M | -/LC | In-Fr | 1 | FI | MC |
| Seiurus aurocapilla | 1 | 1 | M | -/LC | In | 1 | FI | U |
| Setophaga nigrescens | 1 | – | M | -/LC | In | 1 | – | – |
| Setophaga ruticilla | 1 | – | M | -/LC | In | 1 | – | – |
| Tyrannus melancholicus | 1 | – | R | -/LC | In | 1 | VM | C |
| Accipiter striatus | – | 1 | R | Pr/LC | Ca | 3 | – | – |
| Archilochus colubris | – | 1 | M | -/LC | Ne | 1 | FI | UM |
| Table 1. Continued | ||||||||
| Species | # detections by point counts | # captures in mist nets | Migratory status | Protection status | Trophic guild | Size category | Habitat preference | Stratum preference |
| Arremon brunneinucha* | – | 14 | R | -/LC | In-Fr | 2 | FG | M |
| Basileuterus culicivorus | – | 2 | R | -/LC | In | 1 | FG | UM |
| Campylopterus hemileucurus | – | 20 | R | -/LC | Ne | 1 | – | – |
| Cardellina canadensis | – | 1 | M | -/LC | In | 1 | FI | U |
| Catharus aurantiirostris* | – | 2 | R | -/LC | Fr | 1 | FG | U |
| Chlorestes candida | – | 1 | R | -/LC | Ne | 1 | – | – |
| Eugenes fulgens | – | 2 | R | -/LC | Ne | 1 | FE | M |
| Lampornis amethystinus | – | 21 | R | -/LC | Ne | 1 | FG | U |
| Melanotis caerulescens* | – | 1 | R | -/LC | In-Fr | 2 | FE | U |
| Myiodynastes luteiventris* | – | 2 | R | -/LC | In-Fr | 2 | FI | MC |
| Pampa curvipennis | – | 12 | R | -/LC | Ne | 1 | VM | UM |
| Pitangus sulphuratus* | – | 1 | R | -/LC | In-Fr | 2 | FG | A |
| Saucerottia beryllina | – | 12 | R | -/LC | Ne | 1 | – | – |
| Saucerottia cyanocephala | – | 19 | R | -/LC | Ne | 1 | FE | UM |
| Selasphorus rufus | – | 1 | M | -/NT | Ne | 1 | – | – |
| Stelgidopteryx serripennis | – | 1 | R | -/LC | In | 1 | FG | C |
| Vireo gilvus* | – | 1 | R | -/LC | In-Fr | 1 | FI | C |
*Seed dispersing birds: based on field data from Hernández-Dávila et al. (2022) and Hernández-Ladrón De Guevara et al. (2012). Migratory status: from Howell and Webb (1995) and Sibley (2000). Size: from Martínez-Morales (2001), Sibley (2000), and birds captured in mist nets (this study). Habitat preference and stratum preference: from Martínez-Morales (2001).
The most common species recorded visually was Chlorospingus flavopectus (Common Chlorospingus) with 234 detections (Fig. 5A), followed by Cardellina pusilla (Wilson’s Warbler) with 102, Psilorhinus morio (Brown Jay) with 79 and Myadestes occidentalis (Brown-backed Solitaire) with 51 detections (see species detection data in Table 1). These 4 dominant species accounted for 45% of total detections recorded in the point counts. There were 10 species with only 2 detections (i.e., doubletons) and 18 species with only 1 (singletons), and these 28 extremely rare species accounted for less than 5% of total detections and 42% of the 67 species recorded in the point counts. In general, the pattern of dominance by the 4 species mentioned occurred in each riparian corridor sampled, concentrating most of the bird detections at each site, with several species having much fewer detections per site (Fig. 5B).
Bird species richness per site estimated in point counts varied from 9 (TM site) to 33 species (MA site), with 10 of the 14 segments sampled having fewer than 20 species each. Similarity between sites was low (Jaccard index, J < 0.4) for most of the paired comparisons (Table 2), with the highest level of similarity between the TL and AB sites (0.53) and the lowest between MA and AF (0.10).
Discussion
In general, and pooling all sampled riparian corridors, the sample completeness was high (97%), recording a total richness of 86 bird species in the 14 segments sampled in the fragmented cloud forest landscape of central Veracruz, Mexico. This richness is comparable to that reported in similar studies carried out in relatively large (> 3 ha) remnant fragments of cloud forest in the same region, where up to 75-100 bird species have been detected (Rueda-Hernández et al., 2015; Serna-Lagunes et al., 2023). These 14 riparian corridors harbor 24% of the bird richness reported for cloud forest throughout the entire state of Veracruz (Navarro-Sigüenza et al., 2014). The richness recorded in our study represents between 25 and 43% of the avifauna reported in other regions of Mexico with cloud forest, where 196 to 335 bird species have been found (Martínez-Morales, 2007; Navarro-Sigüenza et al., 2014). The richness detected suggests that riparian corridors are important elements in deforested landscapes of cloud forest for numerous birds. Other studies in different sites have shown that these elements of the landscape can serve as refuges, foraging areas and even as reproductive (nesting) sites (Hawes et al., 2008; de Zwaan et al., 2022), as well as making it possible for birds to move across large open areas in agricultural landscapes (Gillies & St. Clair, 2010; Pliscoff et al., 2020). Both resident species and migratory birds, visit and use forested riparian corridors during their autumn-winter stay in the tropics (Skagen et al., 1998; Villaseñor-Gómez, 2008). As shown in our results: 24% of the recorded species were North American migratory species. The migratory bird Cardellina pusilla, abundant in cloud forest as well as in shaded coffee plantations in Veracruz (Navarro-Sigüenza et al., 2014), was the second most common species in our study. By far, the most dominant species in our study was Chlorospingus flavopectus, a resident bird regarded as a forest generalist that is common to old-growth forest, forest edge habitats and patches of secondary forest (Cruz-Angón et al., 2008; Martínez-Morales, 2007; Renner et al., 2006). Myadestes occidentalis was the fourth most dominant species in riparian corridors, which is considered a typical cloud forest species (Caballero-Cruz et al., 2020). This pattern of dominance is similar to that recorded by Martínez-Morales (2001) who reported C. flavopectus, M. occidentalis, Henicorhina leucophrys, Catharus mexicanus, and Trogon mexicanus as the most dominant species in conserved fragments of cloud forest. Except for T. mexicanus, the mentioned species were recorded in this study. Another very common bird in our study was Psilorhinus morio, a habitat generalist associated with disturbed areas, and common in small forest fragments of cloud forest (Serna-Lagunes et al., 2023), in rainforest riparian corridors that cross pastures (Graham et al., 2002), as well as in open agricultural areas with scant arboreal cover (Cerezo et al., 2009). It is important to mention that P. morio is a large species (> 100 gr), only 12% of the species recorded belong to this size category, while 65% are small species (< 40 gr). The conversion of forest for agricultural purposes mainly affects the presence of large bird species due to the reduction in food availability as well as fewer nesting and roosting sites (Gomes et al., 2008; Martínez-Morales, 2001). Thus, large bird species are the most strongly affected by the reduction of forest cover, while small birds are more vagile and tolerant to deforestation, explaining the predominance of species smaller than 40 g in the riparian corridors studied.
Table 2
Similarity distance in bird species composition (Jaccard index) among the 14 riparian corridors sampled (upper-right side of Table) with point counts (see Materials and methods), showing the number of species shared between riparian corridors (lower-left side), and the total number of species in each corridor (diagonal black cells). Gray-shaded cells highlight the highest and lowest values of similarity distance. The names of riparian corridors sampled are abbreviated as in Figure 1.
| AC | AB | AF | TR | GR | MA | ME | MG | PD | RM | TL | TF | TM | VH | |
| AC | 21 | 0.36 | 0.39 | 0.39 | 0.37 | 0.26 | 0.31 | 0.37 | 0.43 | 0.29 | 0.26 | 0.34 | 0.30 | 0.27 |
| AB | 9 | 13 | 0.33 | 0.35 | 0.43 | 0.24 | 0.30 | 0.38 | 0.39 | 0.32 | 0.53 | 0.35 | 0.38 | 0.32 |
| AF | 9 | 6 | 11 | 0.21 | 0.24 | 0.10 | 0.14 | 0.29 | 0.36 | 0.17 | 0.33 | 0.26 | 0.43 | 0.28 |
| TR | 11 | 8 | 5 | 18 | 0.52 | 0.24 | 0.29 | 0.21 | 0.32 | 0.27 | 0.19 | 0.29 | 0.23 | 0.25 |
| GR | 11 | 10 | 6 | 13 | 20 | 0.33 | 0.32 | 0.24 | 0.39 | 0.33 | 0.27 | 0.36 | 0.32 | 0.33 |
| MA | 11 | 9 | 4 | 10 | 13 | 33 | 0.31 | 0.17 | 0.27 | 0.43 | 0.21 | 0.28 | 0.17 | 0.25 |
| ME | 9 | 6 | 3 | 7 | 8 | 11 | 13 | 0.21 | 0.33 | 0.37 | 0.18 | 0.35 | 0.29 | 0.25 |
| MG | 10 | 8 | 6 | 6 | 7 | 7 | 5 | 16 | 0.35 | 0.25 | 0.32 | 0.42 | 0.32 | 0.27 |
| PD | 12 | 9 | 8 | 9 | 11 | 11 | 8 | 9 | 19 | 0.30 | 0.33 | 0.48 | 0.33 | 0.35 |
| RM | 10 | 9 | 5 | 9 | 11 | 17 | 10 | 8 | 8 | 24 | 0.32 | 0.40 | 0.22 | 0.24 |
| TL | 7 | 9 | 6 | 5 | 7 | 8 | 4 | 7 | 7 | 9 | 13 | 0.29 | 0.29 | 0.25 |
| TF | 10 | 8 | 6 | 8 | 10 | 11 | 8 | 10 | 12 | 12 | 7 | 18 | 0.35 | 0.43 |
| TM | 7 | 6 | 6 | 5 | 7 | 6 | 5 | 6 | 8 | 6 | 5 | 7 | 9 | 0.40 |
| VH | 7 | 6 | 5 | 6 | 8 | 9 | 5 | 6 | 10 | 7 | 5 | 9 | 6 | 12 |
Figure 2. Characterization of the avifauna recorded in 14 segments of cloud forest riparian corridors sampled in central Veracruz, Mexico. A, Trophic guild: insectivore (In), frugivore (Fr), nectarivore (Ne), carnivore (Ca), omnivore (Om), and granivore (Gr); B, forest stratum preference: canopy (C), midstory and canopy (MC), midstory (M), understory and midstory (UM), understory (U), and all strata (A); C, habitat preference: forest generalist (FG), forest interior (FI), forest edge (FE), and vegetation matrix (VM); D, size (body mass): small (1 – 40 g), medium (40 – 100 g) and large (100 – 500 g).
Insectivorous birds were the richest and most abundant trophic guild in the riparian corridors sampled and are also the most common guild in intact cloud forest (Martínez-Morales, 2007). Frugivorous birds were the second richest guild and were relatively abundant in our riparian corridors. These species, together with omnivorous species, are particularly important in forest regeneration due to their role as seed dispersers of forest plants. Some of the most important frugivores that are efficient dispersers of cloud forest trees, shrubs and other zoochorous plants include M. occidentalis, C. flavopectus, P. morio, and several species of the genera Turdus, Catharus, and Euphonia (Hernández-Dávila et al., 2022; Hernández-Ladrón De Guevara et al., 2012), all of which were recorded in our study. Seed dispersal by birds is crucial for cloud forest restoration since most plant species native to this forest depend on vertebrate frugivores for dispersal (Jordano et al., 2011). Forest frugivores usually avoid open areas that are devoid of perching sites, and this is one of the strongest limitations to forest restoration due to the limited or absent immigration of woody plant seeds into agricultural areas (Holl et al., 2000). However, it has been recorded that the density of linear forest or wooded patches such as riparian strips and live fences can increase bird diversity in agricultural landscapes (Wilson et al., 2017). In this sense cloud forest riparian corridors could contribute to species movement across the landscape. Thus, forested riparian corridors within the agricultural matrix facilitate that frugivorous birds will visit these disturbed sites and disperse seeds across and into the site. Another group of birds that is crucial for plant reproduction are the nectarivorous species, with several species of hummingbirds being particularly important. Of the 26 hummingbird species reported for Mexican cloud forest (Navarro-Sigüenza et al., 2014), 9 were recorded in the riparian corridors that we studied. The presence of birds that are seed or pollen vectors in riparian corridors contributes greatly to connectivity in anthropic landscapes and are essential to biodiversity conservation and forest regeneration and restoration in these landscapes.
Figure 3. Species accumulation curve and Hill numbers of the bird community recorded in 14 cloud forest riparian corridors. Shaded area delimits 95% confidence intervals.
As expected by the high degree of anthropic disturbance in the landscape we studied, the richest groups of birds in the cloud forest riparian corridors were forest generalist species and those associated with the vegetation of the agricultural matrix (sensu Martínez-Morales, 2001), which together represented 53% of the avifauna we recorded. The presence and wide distribution of different arboreal elements within the current anthropic landscape could explain the presence of forest interior bird species in riparian corridors. Changes in the structure and floristic composition of the original vegetation of a given site, resulting from human activities also lead to changes in the community attributes of the avifauna (Martínez-Morales, 2005). Among the most salient changes in the forested riparian corridors that cross agricultural areas is the high abundance of plants that are common in large canopy gaps or associated with disturbed sites, including several species of the families Piperaceae, Melastomataceae, and Rubiaceae, which were common in our study (Hernández-Dávila et al., 2020). These vegetation changes are more favorable to habitat generalists than to bird species associated with the interior of large forest fragments, explaining the lower proportion of species whose preferred habitat is the forest interior in our results. Studies in different regions report an increase in the richness and abundance of habitat generalist birds in forest edge habitats, where forest interior birds decrease (de Zwaan et al., 2022; Watson et al., 2004; Wilson et al., 2017).
For cloud forest, Martínez-Morales (2005) found that fragment size positively affects the richness and abundance of both generalist and forest-interior birds. In addition, for riparian corridors of cloud forest, Hernández-Davila et al. (2021) found that the richness and abundance of generalist and specialist birds showed a differential response to the amount of forest and urban cover in the vicinity of riparian strips. The percentage of urban cover near the riparian strip negatively affected the abundance of forest interior species and positively affected generalist species, whereas surprisingly the amount of forest area nearby the strip does not seem to influence the richness and abundance of birds using the riparian strip. These results could explain the differences between the number of generalist and interior species found in our study, as well as the differences of the Hill diversity values among the sampled corridors. Although this study did not analyze explicitly aspects of landscape configuration, it is relevant to mention that, although the riparian corridors are narrow remnants of just a few meters wide (< 10 m), both generalist and forest interior species were recorded in them. This suggests that these remnants may harbor a wealth of bird species regardless of their habitat preference. In fact, 2 of the bird species recorded are endemic to Mexico, another 4 are protected by law and 1 is threatened. This highlights the importance of riparian corridors as reservoirs of birds within anthropic landscapes, particularly species native to the cloud forest including resident and migratory birds. It is important to note that 3 of the species recorded in riparian corridors; Quiscalus mexicanus, Rupornis magnirostris, and Pyrocephalus rubinus could be regarded as urban birds (Maya-Elizarrarás, 2011; Ruelas & Aguilar, 2010).
As far as we know, this study is the first to describe and characterize the avifauna present in riparian corridors of the threatened cloud forest, however, it is important to take into account that, despite the fact that our point counts had a sufficient separation and elapsed time between observations to warrant independent detections and no overflying individuals or auditory recordings were included, we might have overestimated species abundances, in particular because birds move frequently along the forest strips (personal observation, OHD). Also, because mist nets capture understory species more frequently, species richness of mid- to high- strata birds are usually underestimated with nets. This study focused on riparian corridors and did not include other types of natural remnants or conserved forest fragments present in the region and part of the current mosaic of the anthropic fragmented landscape. Therefore, more studies are needed to determine the importance of riparian corridors for conserving bird diversity in comparison with other natural remnants of cloud forest.
Figure 4. Species accumulation curve and Hill numbers (q0, q1, and q2) of birds recorded in each riparian corridor. Sites are ordered according to richness; dark blue curves correspond to the riparian corridors with the highest richness (e.g., MA), while light blue curves correspond to the sites with the lowest richness (e.g., AB). Shaded area delimits 95% confidence intervals. Abbreviations of riparian corridors can be found in Figure 1.
Figure 5. A, Rank-abundance curve for the avifauna recorded in 14 segments of cloud forest riparian corridors in central Veracruz, Mexico. The 4 most common species were: C. flavopectus, C. pusilla, P. morio, and M. occidentalis, the number of total detections is given in parenthesis; B, Rank-abundance curve in each riparian corridor. Sites are ordered according to number of detections; dark blue curves correspond to the riparian corridors with the highest number of detections (e.g., LM), while light blue curves correspond to the sites with the lowest number of detections (e.g., AB). Abbreviations of riparian corridors can be found in Figure 1.
In conclusion, our results show that the cloud forest riparian corridors that cross open agricultural areas harbor a diverse assemblage of bird species, which includes not only those tolerant of or associated with disturbance, but also species that are typical of intact patches of old-growth cloud forest. These would be absent in areas that are completely devoid of trees. Additionally, several of the birds recorded in our study are effective seed dispersers of cloud forest plants. For current deforested landscapes, this strongly suggests that cloud forest riparian corridors could be key elements in forest restoration efforts and for the conservation of bird biodiversity in agricultural landscapes. Landscape management plans designed to encourage the permanence and sustainable management of these forested corridors within the agricultural matrix will not only help in the conservation of the avifauna, but also in the restoration of degraded landscapes, thanks to the ecosystem services provided by birds, including the pollination and seed dispersal of plants native to the cloud forest.
Acknowledgement
We are grateful to María de los Ángeles García and Diana Vázquez for their valuable help in the field. The Instituto de Ecología, A.C. and Idea Wild provided the space and equipment that made this study possible. Bianca Delfosse translated the text from the original in Spanish and edited subsequent versions of the manuscript. We also thank two anonymous reviewers for their suggestions to improve the manuscript. This work was supported by The Rufford Foundation (grant number 20471-1 to OHD) and the Consejo Nacional de Ciencia y Tecnología (grant numbers CONACYT-CB-2016-01 to VJS, and graduate scholarship CONACYT-285962 to OHD).
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Diversidad funcional y composición de comunidades de insectos en niveles diferentes de perturbación
Functional diversity and composition of insect communities at different levels of disturbance
Víctor Manuel Caballero-Chan, Alejandra González-Moreno*, Horacio Salomón Ballina-Gómez y Carlos Juan Alvarado-López
Tecnológico Nacional de México, Instituto Tecnológico de Conkal, División de Estudios de Posgrado e Investigación, Av. Tecnológico s/n, 97345 Conkal, Yucatán, México
*Autor para correspondencia: alejandra.gonzalez@itconkal.edu.mx (A. González-Moreno)
Recibido: 18 enero 2025; aceptado: 30 mayo 2025
Resumen
La diminución en la cobertura vegetal y la perturbación antropogénica tienen efectos negativos sobre la diversidad de insectos. Este trabajo tuvo como objetivo evaluar la diversidad funcional y la composición de comunidades de insectos fitófagos y benéficos en diferentes niveles de perturbación de Yucatán, México. Se instalaron 6 trampas Malaise por sitio, durante 5 meses en temporada de lluvias. Los ejemplares se identificaron a nivel familia y grupo funcional; se analizó la diversidad en términos de riqueza, familias comunes y dominantes de cada grupo funcional. Se registraron 25,872 individuos de 106 familias, 12 órdenes y 4 grupos funcionales (fitófagos, polinizadores, depredadores y parasitoides). Aunque la riqueza de familias fue similar, la diversidad de familias comunes y dominantes mostró diferencias en los sitios con niveles medios y altos de perturbación. Estos resultados sugieren que algunas familias son exitosas en niveles altos de perturbación y otras disminuyen su diversidad. Las familias dominantes de los fitófagos, polinizadores, parasitoides y depredadores fueron: Pyralidae, Geometridae, Tachinidae y Coccinellidae, respectivamente.
Palabras clave: Urbanización; Grupos funcionales; Fitófagos; Insectos benéficos
Abstract
Decreased vegetation cover and anthropogenic disturbance can negatively impact insect diversity. This research aimed to assess the functional diversity and community composition of phytophagous and beneficial insects at different levels of disturbance in Yucatán, Mexico. Six Malaise traps were deployed at each site during 5 months in the rainy season. Specimens were identified to the family and functional group levels, and diversity was analyzed based on family richness, as well as the composition of common and dominant families for every functional group. A total of 25,872 individuals representing 106 families, 12 orders, and 4 functional groups were recorded (phytophagous, pollinators, predators and parasitoids). While family richness was comparable across sites, the diversity of common and dominant families differed between areas with medium and high levels of disturbance. These results suggest that some families thrive under high disturbance levels, whereas other experiences a decline in diversity. The dominant families of phytophagous, pollinators, parasitoids and predators were: Pyralidae, Geometridae, Tachinidae, and Coccinellidae, respectively.
Keywords: Urbanization; Functional groups; Phytophagous; Beneficial insects
Introducción
En México, un país considerado megadiverso, el crecimiento de las zonas urbanas y la intensificación de la agricultura, están poniendo en riesgo la diversidad de insectos nativos presentes en zonas naturales (Martínez-Ramos et al., 2016), así como ocasionando cambios en el comportamiento de distintos grupos de insectos, que comprometen funciones ecológicas críticas como la polinización, el control biológico de plagas y la descomposición de la materia orgánica (Wagner et al., 2021). La perturbación antropogénica derivada de la urbanización, que incluye la construcción de infraestructura y el desarrollo de asentamientos humanos, junto con la agricultura intensiva, la deforestación, la fragmentación del hábitat, la contaminación y cambio climático, está generando una presión sin precedentes sobre la biodiversidad (Betts et al., 2019; Fahrig et al., 2019). La disminución de especies de vertebrados, como aves, anfibios y mamíferos está bien documentada; sin embargo, la pérdida de diversidad en invertebrados es menos conocida y se desconoce si está ocurriendo a la misma velocidad que en otros grupos; aunque estudios más recientes, incluidos varios metaanálisis, han evidenciado el declive de los insectos (Wagner et al., 2021).
En particular, se ha demostrado que la urbanización y agricultura intensiva afectan negativamente a los polinizadores, principalmente por la pérdida de hábitat, cambios en la disponibilidad de alimentos y la alteración de sus patrones de comportamiento (Biella et al., 2022; Fisogni et al., 2020; Tavares-Brancher et al., 2024). Pero, la urbanización no solo afecta a los polinizadores, sino que en general contribuye a la disminución de insectos debido a múltiples factores, como el incremento en la intensidad lumínica (Boyes et al., 2021), en la temperatura, evaporación y la exposición al viento y a diversos contaminantes, que alteran su comportamiento y actividad (Dirzo, 2014; Wagner et al., 2021); especialmente aquellos que ocupan altos niveles tróficos como los parasitoides y depredadores (Betancourt et al., 2021; Janzen y Hallwachs, 2021). Esta pérdida de diversidad no solo es preocupante por la pérdida de especies en sí misma, sino también por la disminución de las múltiples funciones ecológicas que realizan los insectos (Fenoglio et al., 2020; Wagner et al., 2021).
En México se han realizado algunos estudios sobre cómo la diversidad de insectos varía en varios niveles de perturbación, con resultados contrastantes, dependiendo del taxón a estudiar, del grupo funcional al que pertenecen y de las características de los sitios perturbados. En general, se ha demostrado que la urbanización tiene un impacto negativo sobre la diversidad de abejas (Muñoz-Urías et al., 2025), coleópteros (Cortés-Arzola y León-Cortés, 2021) y hormigas, principalmente de hábitos arborícolas (Roche-Ortega y Castaño- Meneses, 2015) y mariposas (Ramírez-Restrepo y Halffter, 2013). Actualmente, la península de Yucatán está siendo amenazada por el crecimiento urbano descontrolado y las prácticas agrícolas no sostenibles, por lo que se planteó la siguiente pregunta de investigación ¿cómo varían las comunidades de insectos en términos de diversidad funcional y composición, en diferentes niveles de perturbación? Tomando en cuenta que los resultados de las evaluaciones de diversidad pueden variar según el nivel taxonómico considerado y los requerimientos ecológicos del taxón a estudiar (Fenoglio et al., 2020), en este trabajo se planteó abordarlo a nivel de familia, considerando el grupo funcional que conforman y la función que llevan a cabo como fitófagos, depredadores, parasitoides o polinizadores.
El objetivo del presente trabajo fue evaluar la diversidad funcional y la composición de comunidades de insectos fitófagos y benéficos en diferentes niveles de perturbación de Yucatán; para ello se plantearon las siguientes hipótesis: la diversidad de insectos fitófagos será mayor en zonas con mayor nivel de perturbación; por el contrario, los insectos benéficos serán más diversos en niveles de perturbación menores y la familia dominante de cada grupo funcional será diferente en cada nivel de perturbación. Realizar estas evaluaciones podría ser significativo para implementar estrategias de conservación dentro de las ciudades, además de que las evaluaciones de la diversidad de invertebrados son prioritarias para avanzar en el entendimiento de la defaunación, principalmente en regiones tropicales (Dirzo, 2014; Wagner et al., 2021).
Materiales y métodos
El estudio se llevó a cabo en 3 zonas de Yucatán que se eligieron considerando el grado de perturbación, de acuerdo con el índice de disturbio (ID), según el método propuesto por Martorell y Peters (2005). En cada zona, se seleccionaron 2 sitios al azar, se utilizaron fotografías aéreas (INEGI SINFA 1:250 000, 2019 CLAVE F16-10 LÍNEA 166) que fueron importadas al programa Arc view 3.1. El ID se basa en la cuantificación de 15 parámetros, los cuales están comprendidos en 1 de 3 categorías: 1) cría de ganado (frecuencia de excrementos de cabra, frecuencia de excremento de vaca, ramoneo, caminos para el ganado y compactación del suelo); 2) actividades humanas (extracción de leña, número de caminos, superficie de senderos, proximidad de asentamientos humanos, cercanía a núcleos de actividad humana, porcentaje de uso del suelo y evidencia de incendios forestales) y 3) degradación del suelo (porcentaje de erosión, presencia de islas de erosión y superficie totalmente modificada). Una vez que se estimó el índice de perturbación en cada sitio, se establecieron 3 niveles de perturbación: alto (índice de perturbación entre 10 y 15), medio (entre 2.1 y 5) y bajo (entre 0 y 2). Así, se seleccionaron 2 sitios por nivel de afectación, resultando 6 sitios en total (tabla 1).
La zona con alto nivel de perturbación, a la cual se denominó zona urbana, se ubicó alrededores de la ciudad de Mérida (21°01’ N, 89°33’ O), al norte del estado, dominada por un paisaje urbano, con 75% de cobertura gris y apenas 10% de cobertura de material vegetal (Biles y Lemberg, 2023). La zona con nivel medio de perturbación, nombrada zona periurbana, se localiza en el municipio de Conkal (21°05’ N, 89°30’ O), al este del municipio de Mérida, con 40% de cobertura gris y 35% de cobertura vegetal, dominada por cultivos de maíz y vegetación circundante de selva baja caducifolia (Rzedowski, 1978). La zona, con bajo nivel de perturbación, llamada zona natural, se localiza en la reserva privada “Komchén de los Pájaros” (21°13’ N, 89°19’ O), al oeste de los municipios de Dzemul y Telchac, presenta una cobertura vegetal de 80% y una cobertura gris inferior a 5% (fig. 1); la vegetación es selva baja caducifolia, conformada principalmente por varias especies de la familia Fabaceae: Piscidia piscipula (jabín), Caesalpinia gaumeri (kitinché), Lysiloma latisiliquum (tzalam) y la familia Burseraceae: Bursera simaruba (palo mulato) (Flores y Espejel, 1994). Las 3 zonas presentan un clima cálido subhúmedo con lluvias en verano con rangos de precipitación anual que van de 1,050 mm a 1,200 mm. La temperatura media anual es de 26 a 28 °C.
Tabla 1
Índice y categoría de perturbación de los 6 sitios seleccionados en México.
| Nombre del sitio | Tipo de sitio | Índice | Nivel |
| Mérida 1 | Huerto en ciudad | 10.33 | Alto |
| Mérida 2 | Huerto en ciudad | 11.35 | Alto |
| Conkal 1 | Huerto periurbano con cultivo de maíz | 2.22 | Medio |
| Conkal 2 | Huerto periurbano con cultivo de maíz | 3.59 | Medio |
| Komchén de los Pájaros 1 | Selva baja caducifolia | 0.78 | Bajo |
| Komchén de los Pájaros 2 | Selva baja caducifolia | 0.81 | Bajo |
Se llevaron a cabo muestreos sistemáticos durante la temporada de lluvias, de julio a diciembre de 2023. En cada nivel de perturbación se instalaron 2 trampas de intercepción tipo Malaise, separadas por más de 2 km, para asegurar la independencia de las muestras; estas trampas fueron seleccionadas por su alta eficacia en la captura pasiva de insectos voladores, incluidos lepidópteros diurnos y nocturnos (Schmidt et al., 2019), además de ser uno de los métodos de muestreo mayormente utilizados para hacer evaluaciones de biodiversidad, y permiten obtener un muestreo representativo de los ensambles presentes en cada sitio (Chan-Canché et al., 2020; Kaczmarek et al., 2022). En los sitios urbanos las trampas fueron colocadas en jardines de casas particulares; en la zona periurbana se colocaron en cultivos de maíz y en la zona natural en parches de vegetación. Las trampas funcionaron ininterrumpidamente durante 5 meses con cortes quincenales de recolecta, resultando un total de 10 muestras por sitio. Los botes recolectores de las trampas tenían 1 L de etanol desnaturalizado al 70%, el cual era reemplazado en cada recolecta. Cada muestra se procesó según las técnicas curatoriales convencionales, en el laboratorio de plagas agrícolas del Instituto Tecnológico de Conkal; todos los insectos se conservaron en etanol al 70% y se seleccionaron algunos ejemplares para su montaje en seco en alfileres entomológicos para su posterior identificación; en el caso de los lepidópteros, se empleó la técnica de relajación en cámara húmeda, extensión de alas sobre planchas entomológicas y posterior secado para su correcta preservación y manejo en la colección. Posteriormente se realizó la identificación taxonómica de los ejemplares a categoría de familia, utilizando claves especializadas en insectos de Latinoamérica, como las de Goulet y Huber (1993), Borror y White (1998), Arnett (2000), Triplehorn y Johnson (2005).
Se analizó la representatividad del muestreo con el software EstimateS 9.10 mediante curvas de acumulación, con el estimador no paramétrico jackknife 1, conocido por ser uno de los estimadores menos sesgados para muestras pequeñas (Magurran, 2004); se utilizaron las 10 fechas de recolecta como medida del esfuerzo de muestreo, con un total de 3,600 horas de recolecta por trampa. Asimismo, se analizaron las diferencias entre la riqueza de familias entre sitios, considerando los intervalos de confianza al 95% de 1,000 remuestreos calculados mediante la prueba de bootstrap (Colwell y Elsensohn, 2014); la no superposición de los intervalos de confianza indica diferencias estadísticamente significativas (Colwell, 2006). Para realizar el análisis de diversidad de insectos en cada nivel de perturbación, las familias identificadas se agruparon considerando el grupo funcional que conforman: fitófagos, polinizadores, parasitoides o depredadores; para cada grupo el análisis de diversidad fue calculado mediante medidas de diversidad verdadera, usando el software SPADE (Chao y Shen, 2010). Estas medidas contemplan 3 niveles de diversidad basadas en los números de Hill, qD (Jost, 2006): 0D, se refiere a la riqueza de familias solamente; 1D, es la diversidad ecológica si todas las familias tuvieran la misma importancia relativa, usa el inverso del exponencial de la entropía de Shannon; y 2D que considera solo a las familias dominantes, mediante el inverso del índice de Simpson (Moreno et al., 2011). Todos los valores de qD se calcularon por separado y se tomó en cuenta cada zona de manera individual (zona urbana, zona periurbana y zona natural), lo que permitió evaluar la diversidad por cada zona de perturbación para después comparar entre sitios, usando intervalos de confianza de 95% que fueron calculados mediante la prueba de bootstrap, para saber si existen diferencias significativas entre zonas, la no superposición de los intervalos de confianza indica diferencias estadísticamente significativas (Colwell, 2006).
Posteriormente, se analizó la composición de los ensambles de cada grupo funcional en términos de la distribución de abundancias de cada familia, para lo que se construyeron curvas de rango-abundancia para cada nivel de perturbación (Whittaker, 1972).
Resultados
Se recolectaron 25,872 individuos pertenecientes a 4 grupos funcionales: fitófagos, polinizadores, parasitoides y depredadores, clasificados en 12 órdenes y 106 familias (tabla 2). Las curvas de acumulación indican que el muestreo tuvo una eficiencia de 83.4%, (familias observadas = 86; jackknife 1 = 103.1) para la zona urbana, de 84.7% para la zona periurbana (familias observadas = 90; jackknife 1 = 106.2) y de 83.5 % para la zona natural (familias observadas = 73; jackknife 1 = 87.4), con menor riqueza para la zona natural, con intervalos de confianza al 95% (fig. 2).
En el grupo de los fitófagos, la riqueza de familias (0D) y la diversidad de familias dominantes (2D) no mostraron diferencias significativas entre zonas, considerando el sobrelapamiento de los intervalos de confianza al 95%; únicamente la diversidad de familias comunes (1D) muestra que la zona periurbana tiene el valor más alto (tabla 3). La familia Pyralidae se registró dominando en las zonas periurbana y natural, en cambio, la familia Cicadellidae, lo hizo únicamente en la zona urbana (fig. 3A).
Los insectos polinizadores, al igual que los fitófagos, presentaron diferencias en la diversidad de familias comunes (1D), siendo los más diversos en la zona urbana con los intervalos de confianza al 95%. La riqueza (0D) y diversidad de familias dominantes (2D) fueron similares en los 3 sitios (tabla 3). Sin embargo, pese a que la diversidad fue similar, las familias dominantes fueron diferentes en cada nivel de perturbación: Geometridae en los sitios periurbanos con maíz, Nymphalidae en el sitio natural y Erebidae en la zona urbana (fig. 3B).
Tabla 2
Grupos funcionales, órdenes y familias de insectos identificados en un gradiente de perturbación de Yucatán, México.
| Grupo | Orden | Familia | Zona urbana | Zona periurbana | Zona natural | Total |
| Fitófago | Lepidoptera | Pyralidae | 2,298 | 2,117 | 1,459 | 5,874 |
| Crambidae | 219 | 969 | 505 | 1,693 | ||
| Tortricidae | 242 | 363 | 164 | 769 | ||
| Noctuidae | 3 | 286 | 23 | 312 | ||
| Tineidae | 12 | 116 | 54 | 182 | ||
| Oecophoridae | 0 | 1 | 0 | 1 | ||
| Hemiptera | Cicadellidae | 3,063 | 1,034 | 281 | 4,378 | |
| Diaspididae | 309 | 24 | 4 | 337 | ||
| Psyllidae | 1 | 122 | 2 | 125 | ||
| Aphididae | 43 | 11 | 19 | 73 | ||
| Derbidae | 9 | 25 | 3 | 37 | ||
| Cicadidae | 13 | 8 | 10 | 31 | ||
| Cixiidae | 9 | 7 | 7 | 23 | ||
| Liviidae | 19 | 0 | 1 | 20 | ||
| Rhyparochromidae | 1 | 18 | 0 | 19 | ||
| Berytidae | 7 | 3 | 0 | 10 | ||
| Delphacidae | 2 | 6 | 1 | 9 | ||
| Alydidae | 3 | 4 | 0 | 7 | ||
| Membracidae | 3 | 1 | 1 | 5 | ||
| Cercopidae | 4 | 0 | 0 | 4 | ||
| Coreidae | 0 | 4 | 0 | 4 | ||
| Miridae | 0 | 2 | 1 | 3 | ||
| Pentatomidae | 1 | 2 | 0 | 3 | ||
| Pyrrhocoridae | 0 | 3 | 0 | 3 | ||
| Lygaeidae | 0 | 2 | 0 | 2 | ||
| Triozidae | 0 | 0 | 2 | 2 |
| Tropiduchidae | 1 | 1 | 0 | 2 | ||
| Cydnidae | 1 | 0 | 0 | 1 | ||
| Dictyopharidae | 0 | 1 | 0 | 1 | ||
| Rhopalidae | 0 | 1 | 0 | 1 | ||
| Tingidae | 1 | 0 | 0 | 1 | ||
| Diptera | Drosophilidae | 496 | 284 | 98 | 878 | |
| Ulidiidae | 158 | 71 | 47 | 276 | ||
| Tephritidae | 1 | 11 | 8 | 20 | ||
| Psilidae | 1 | 0 | 0 | 1 | ||
| Coleoptera | Chrysomelidae | 48 | 62 | 34 | 144 | |
| Curculionidae | 20 | 19 | 14 | 53 | ||
| Mordellidae | 9 | 1 | 3 | 13 | ||
| Bruchidae | 0 | 2 | 0 | 2 | ||
| Orthoptera | Gryllidae | 5 | 15 | 57 | 77 | |
| Acrididae | 1 | 22 | 6 | 29 | ||
| Tettigoniidae | 1 | 6 | 2 | 9 | ||
| Thysanoptera | Phlaeothripidae | 2 | 13 | 0 | 15 | |
| Thripidae | 1 | 3 | 3 | 7 | ||
| Polinizador | Lepidoptera | Geometridae | 187 | 1,672 | 222 | 2,081 |
| Erebidae | 251 | 1,214 | 390 | 1,855 | ||
| Nymphalidae | 25 | 257 | 1,127 | 1,409 | ||
| Pieridae | 77 | 122 | 119 | 318 | ||
| Hesperiidae | 31 | 82 | 58 | 171 | ||
| Pterophoridae | 12 | 38 | 64 | 114 | ||
| Lycaenidae | 23 | 66 | 14 | 103 | ||
| Riodinidae | 2 | 61 | 0 | 63 | ||
| Sphingidae | 14 | 8 | 1 | 23 | ||
| Diptera | Bombyliidae | 25 | 459 | 66 | 550 | |
| Syrphidae | 25 | 103 | 7 | 135 | ||
| Tipulidae | 6 | 25 | 25 | 56 | ||
| Hymenoptera | Apidae | 25 | 14 | 7 | 46 | |
| Halictidae | 1 | 0 | 0 | 1 | ||
| Parasitoide | Diptera | Tachinidae | 106 | 593 | 158 | 857 |
| Pipunculidae | 13 | 18 | 15 | 46 | ||
| Hymenoptera | Braconidae | 146 | 99 | 78 | 323 | |
| Ichneumonidae | 37 | 114 | 63 | 214 | ||
| Encyrtidae | 114 | 28 | 14 | 156 | ||
| Figitidae | 67 | 2 | 2 | 71 | ||
| Chalcididae | 21 | 23 | 3 | 47 | ||
| Bethylidae | 31 | 4 | 11 | 46 | ||
| Eulophidae | 22 | 5 | 6 | 33 | ||
| Eupelmidae | 14 | 5 | 5 | 24 | ||
| Diapriidae | 12 | 4 | 5 | 21 | ||
| Aphelinidae | 12 | 5 | 2 | 19 | ||
| Pteromalidae | 12 | 4 | 3 | 19 | ||
| Mymaridae | 3 | 0 | 11 | 14 | ||
| Eurytomidae | 9 | 3 | 0 | 12 | ||
| Evaniidae | 5 | 2 | 3 | 10 | ||
| Tiphiidae | 4 | 4 | 1 | 9 | ||
| Trichogrammatidae | 5 | 0 | 3 | 8 | ||
| Perilampidae | 4 | 1 | 1 | 6 | ||
| Chrysididae | 0 | 5 | 0 | 5 | ||
| Platygastridae | 3 | 1 | 0 | 4 | ||
| Eucharitidae | 3 | 0 | 0 | 3 | ||
| Tetracampidae | 0 | 1 | 2 | 3 | ||
| Mutillidae | 0 | 2 | 0 | 2 | ||
| Dryinidae | 0 | 0 | 1 | 1 | ||
| Rhopalosomatidae | 0 | 1 | 0 | 1 | ||
| Depredador | Coleoptera | Coccinellidae | 451 | 118 | 51 | 620 |
| Dytiscidae | 5 | 2 | 1 | 8 | ||
| Carabidae | 2 | 3 | 0 | 5 | ||
| Diptera | Dolichopodidae | 250 | 41 | 25 | 316 | |
| Asilidae | 43 | 140 | 59 | 242 | ||
| Scenopinidae | 0 | 4 | 0 | 4 | ||
| Hemiptera | Anthocoridae | 2 | 2 | 1 | 5 | |
| Reduviidae | 1 | 3 | 0 | 4 | ||
| Stenocephalidae | 1 | 0 | 0 | 1 | ||
| Hymenoptera | Crabronidae | 58 | 16 | 18 | 92 | |
| Vespidae | 20 | 22 | 4 | 46 | ||
| Pompilidae | 5 | 8 | 2 | 15 | ||
| Sphecidae | 3 | 2 | 0 | 5 | ||
| Sapygidae | 0 | 2 | 1 | 3 | ||
| Thynnidae | 0 | 2 | 0 | 2 | ||
| Scoliidae | 0 | 0 | 1 | 1 | ||
| Mantodea | Mantidae | 9 | 17 | 8 | 34 | |
| Mantoididae | 2 | 1 | 0 | 3 | ||
| Mecoptera | Bittacidae | 1 | 0 | 0 | 1 | |
| Neuroptera | Chrysopidae | 24 | 73 | 31 | 128 | |
| Berothidae | 0 | 0 | 1 | 1 | ||
| Myrmeleontidae | 0 | 0 | 1 | 1 |
Los parasitoides, mostraron mayores diferencias en términos de diversidad, siendo la zona urbana la que tuvo la mayor diversidad de especies comunes (1D) y dominantes (2D), pese que la riqueza de familias (0D) fue similar en los 3 niveles de perturbación, según los valores de los intervalos de confianza al 95% (tabla 3). La familia Tachinidae, contrario a lo esperado, se registró como dominante en la zona periurbana y natural, pero en la zona urbana, la familia Braconidae fue la dominante (fig. 3C).
Los insectos depredadores presentaron la misma riqueza de familias (0D) en los diferentes sitios, pero fueron más diversos (1D y 2D) en la zona natural y periurbana, de acuerdo con los intervalos de confianza al 95% (tabla 3) con la familia Asilidae dominando las comunidades más diversas; por el contrario, Coccinellidae dominó las comunidades de la zona urbana (fig. 3D).
Discusión
Los resultados de este trabajo fueron contrarios a nuestra hipótesis de investigación, ya que ésta sugiere que en zonas con altos niveles de perturbación como las ciudades, habría menor diversidad de insectos, principalmente especialistas como los parasitoides, considerando la hipótesis del aumento del disturbio, que indica una disminución en la riqueza de artrópodos, particularmente especialistas, conforme aumenta el grado de urbanización (Gray, 1989, en Fenoglio et al., 2020) y la teoría sobre la complejidad estructural de la vegetación, que afirma que a mayor cobertura vegetal en ecosistemas, habrá un mayor número de plantas disponibles, lo que permitirá alojar mayor diversidad de fitófagos y, por consiguiente, de depredadores y parasitoides (González-Moreno et al., 2023; Guo et al., 2021; Neal et al., 2024).
Tabla 3
Diversidad verdadera de grupos funcionales de insectos, como número efectivo de familias, con números de Hill para estimar la riqueza de familias (0D), la diversidad de familias comunes (1D) y la diversidad de familias dominantes (2D) en 3 zonas con diferente nivel de perturbación de Yucatán. *Los intervalos de confianza (IDC) al 95%, indican diferencias significativas.
| Grupos funcionales | Índices de diversidad verdadera | ||
| 0D (95% IDC) | 1D (95% IDC) | 2D (95% IDC) | |
| Fitófagos | |||
| Zona urbana | 50.6 (39.9, 84.8) | 4.8 (4.6, 4.9)* | 3.2 (2.4, 4.1) |
| Zona periurbana | 42.4 (39.0, 56.7) | 7.2 (6.9, 7.4)* | 4.7 (4.0, 5.3) |
| Zona natural | 30.3 (27.7, 43.1) | 5.3 (5.0, 5.6)* | 3.1 (2.5, 3.8) |
| Polinizadores | |||
| Zona urbana | 14.9 (14.1, 25.2) | 6.7 (6.2, 7.2)* | 4.6 (4.2, 5.0) |
| Zona periurbana | 13.0 (13.0, 13.0) | 5.2 (5.0, 5.4)* | 3.7 (3.1, 4.3) |
| Zona natural | 12.5 (12.0, 20.0) | 4.4 (4.2, 4.7)* | 2.9 (2.3, 3.6) |
| Parasitoides | |||
| Zona urbana | 21.0 (21.0, 21.0) | 10.8 (10.0, 11.6)* | 7.7 (7.4, 8.0)* |
| Zona periurbana | 23.6 (22.3, 32.0) | 4.0 (3.6, 4.3)* | 2.3 (1.5, 3.1)* |
| Zona natural | 21.3 (20.2, 29.3) | 6.9 (6.0, 7.8)* | 4.2 (3.6, 4.8)* |
| Depredadores | |||
| Zona urbana | 18.7 (16.4, 31.7) | 4.1 (3.8, 4.5)* | 2.8 (2.0, 3.6)* |
| Zona periurbana | 17.5 (17.0, 22.7) | 7.0 (6.3, 7.7) | 5.0 (4.5, 5.6) |
Las diferencias encontradas en la riqueza de familias de insectos, sin tomar en cuenta su función en los ecosistemas, puede explicarse según la hipótesis de la perturbación media, que propone que la diversidad puede ser mayor en sitios donde la perturbación no es muy frecuente ni muy intensa, comparada con sitios no perturbados o con perturbación intensa (Connell, 1978).
El efecto de la perturbación y heterogeneidad del paisaje sobre la diversidad puede ser diferente dependiendo de la escala de estudio (Corcos et al., 2019), del taxón y del grupo funcional (Fenoglio et al., 2020), como se pudo observar en este trabajo. Las comunidades más diversas de fitófagos se presentaron en los sitios periurbanos con cultivos de maíz, probablemente, por la oferta mayor de alimento que puede representar el cultivo, facilitando el acceso a recursos alimenticios y refugio, que a su vez, favorece una mayor equidad de familias comunes (Landry et al., 2020); además, si consideramos nuevamente la hipótesis del disturbio medio, los sitios periurbanos con un nivel medio de perturbación, estarían alojando mayor diversidad, en este caso de fitófagos. Asimismo, se ha demostrado que zonas perturbadas que integran espacios verdes como parques urbanos, jardines residenciales, huertos verticales y familiares, parcelas de policultivos y parches de vegetación natural, favorecen la diversidad de fitófagos, al proporcionarles recursos alimenticios y refugios suficientes, que les permite adaptarse y prosperar en hábitats alterados por la actividad humana (Landry et al., 2020; Ruiz-Montoya et al., 2014). En los ensambles, la familia Pyralidae fue dominante en las zonas periurbana y natural, debido a su capacidad para aprovechar tanto plantas cultivadas como nativas, lo que le permite alimentarse y completar su ciclo de vida con eficacia (Cepeda, 2017). En cambio, las comunidades de la zona urbana estuvieron dominadas por la familia Cicadellidae, lo que refleja su adaptación a ambientes con altos niveles de perturbación (Trivellone et al., 2021).
La mayor diversidad de polinizadores en ciudad puede estar relacionada con la variabilidad y abundancia de recursos florales de los jardines, parques y áreas verdes presentes en los sitios, al proporcionar ciertas fuentes polínicas que favorecen dicha diversidad; estos resultados son contrarios a la que esperábamos si se considera la teoría sobre el espectro de polinización, que afirma, que a mayor cobertura vegetal en los ecosistemas, habrá un mayor número de plantas con flores, lo que permitirá alojar mayor diversidad de polinizadores. Sin embargo, es importante señalar que esta diversidad estuvo representada por familias de lepidópteros y no por abejas, debido probablemente a que estas últimas son de los grupos más sensibles a la contaminación de las ciudades (Roguz et al., 2023). Aunque se ha demostrado que en áreas urbanas, cuando se crean nuevos hábitats o refugios, como hoteles para polinizadores, jardines florales y corredores de flores silvestres, la diversidad de polinizadores tiende a incrementarse (Francini et al., 2022; Persson et al., 2023); pero el grupo de polinizadores varía dependiendo de ciertos factores asociados a la urbanización, por ejemplo, los abejorros son muy sensibles a la contaminación de las ciudades (Roguz et al., 2023); pero otros grupos pueden tener la capacidad de alimentarse en entornos urbanos (McLeod et al., 2021), como las mariposas (Lepidoptera: Papilionoidea) que son más tolerantes a la urbanización que otros polinizadores como sírfidos, moscas abejorros (Diptera: Syrphidae, Bombyliidae) y abejas (Hymenoptera: Apoidea) (Ávalos-Hernández et al., 2024). Es importante señalar que la diversidad observada es de familias consideradas comunes, lo que está reflejando la adaptación de ciertos polinizadores generalistas, que se adaptan rápidamente a las condiciones altamente perturbadas y son capaces de aprovechar la oferta floral de estos sitios (Deguines et al., 2016; Neumann et al., 2024). Las familias dominantes encontradas en este trabajo se han registrado interactuando con varias especies de plantas en jardines urbanos, realizando la función de polinización (Wonderlin et al., 2019). Geometridae fue dominante en los huertos periurbanos con maíz, probablemente porque son lepidópteros que disminuyen en sitios más urbanizados (Gaona et al., 2021); Nymphalidae fue dominante en la vegetación natural y Erebidae en la ciudad; esto es relevante porque las áreas urbanas y agrícolas alteran los patrones de comportamiento de los lepidópteros nocturnos debido a la luz artificial, mientras que los lepidópteros diurnos en áreas naturales reflejan su dependencia de hábitats con bajo impacto humano (Seymoure, 2018).
La diversidad de parasitoides encontrada, también fue contraria a los patrones esperados, ya que, a mayor cobertura vegetal en los ecosistemas, habrá un mayor número de hospederos disponibles, lo que permitirá alojar mayor diversidad de parasitoides (Parsons y Frank, 2019). Esto puede explicarse por la denso-dependencia de los parasitoides a sus hospederos, incluso en ciudades (Rocha y Fellowes, 2018), probablemente, porque las trampas se colocaron en jardines de casas particulares, los cuales tenían diferentes especies vegetales, ornamentales, frutales y arbustivas que ofrecían microhábitats y recursos alimenticios a los hospederos y por consiguiente a sus parasitoides (Klaus et al., 2024; Lucatero et al., 2024; Start et al., 2020). Otra razón de los valores elevados de diversidad en la zona urbana, puede explicarse por la habilidad de dispersión de los bracónidos, ya que se ha comprobado que los artrópodos que ocupan altos niveles tróficos como depredadores y parasitoides, serán exitosos en sitios urbanos, si tienen alta capacidad de dispersión (Korányi et al., 2022). Tachinidae, que no ha sido registrada como una familia particularmente abundante para la región, fue dominante en los huertos periurbanos con maíz y la vegetación natural; además no es una familia de parasitoides hiperdiversa en comparación con los himenópteros parasitoides, por lo que es menos probable de encontrarse como dominante en los sitios (Kankonda et al., 2018). Por otra parte, se ha registrado que la abundancia de esta familia disminuye en áreas con altas densidades de edificios y calles que actúan como barreras que dificultan la dispersión de los individuos, así como la localización de sus hospederos (Corcos et al., 2019). Por el contrario, la familia Braconidae dominó los sitios de zonas urbanas, lo que podría sugerir su adaptación a ambientes altamente perturbados (Koptur et al., 2024), además de que se ha demostrado que la urbanización favorece especies generalistas capaces de explotar diferentes recursos, como podría ser el caso de Braconidae, parasitoides con mayor variedad en estrategias de desarrollo y biología.
Los depredadores fueron más diversos en áreas periurbanas con maíz y selva, al contrario de los parasitoides, que fueron más diversos en ciudad, probablemente porque al ocupar nichos similares, estén evitando la competencia. Cabe destacar que los depredadores, estuvieron representados mayoritariamente por diferentes familias del orden Coleoptera, que es uno de los grupos de insectos más afectados por la urbanización en términos de riqueza de especies, pero no de abundancia (Fenoglio et al., 2020). Aunque los coleópteros depredadores pueden aprovechar más eficientemente la oferta extendida del alimento que otros enemigos naturales, al adaptarse mejor a los cambios ambientales por urbanización (Gardiner et al., 2021; Liere y Cowal, 2024); como se observó en nuestros resultados, Coccinellidae fue dominante únicamente en la zona urbana, lo que confirma que son los depredadores mejor adaptados a la urbanización, por su capacidad para prosperar en las condiciones microclimáticas características de este ambiente (Kawakami et al., 2016; Meseguer et al., 2024). En cambio, Asilidae fue dominante en las zonas periurbana con maíz y natural, lo que puede reflejar su mayor supervivencia en zonas menos modificadas por la actividad humana (Pascacio-Villafán y Cohen, 2023).
Las diferencias de diversidad observadas para los diferentes ensambles de insectos, reflejan que las zonas urbanas pueden inducir cambios en las comunidades de insectos, e incluso, en algunos casos, favorecer esta diversidad. Pero, esto solo funcionará para cierto grupo de organismos que tengan estrategias tipo “r”, con hábitos generalistas y alta capacidad de dispersión y adaptación, que los hace exitosos en ambientes urbanos (Martinson y Raupp, 2013), como ciertas especies de avispas que pueden ser resistentes a la urbanización (Christie y Hochuli, 2009), como en nuestros resultados representadas por Braconidae; también familias que estén mejor adaptadas a ambientes con altos niveles de perturbación y que pueden colonizar nuevos hábitats creados por la actividad humana y volverse dominantes dentro de las comunidades al verse favorecidas por factores ambientales como la temperatura y la humedad (Adams et al., 2020; Sire et al., 2022). Sin embargo, los resultados contrastantes reportados en la literatura sobre los efectos de la urbanización sobre las comunidades de artrópodos (Arnold, 2022), sugieren que es prioritario continuar con esta línea de investigación.
En conclusión, la riqueza de insectos en sitios con distintos niveles de perturbación fue similar; sin embargo, se encontraron diferencias en la diversidad de familias comunes, esto resulta relevante debido a que los análisis se realizaron a un nivel taxonómico alto, de familia, donde normalmente es difícil detectar variaciones significativas en diversidad. Cabe destacar, que las diferencias fueron con las familias consideradas comunes, que probablemente sean las que presentan biologías generalistas que les permite adaptarse a condiciones adversas como las que existen en zonas perturbadas.
Contrario a lo esperado, la mayor diversidad de insectos no se encontró en la vegetación natural, sino que fue en los sitios periurbanos con cultivos de maíz, con niveles de perturbación moderados, con la familia Pyralidae dominando las comunidades. En la zona urbana, con el mayor grado de perturbación, los grupos de insectos más diversos fueron parasitoides y polinizadores con las familias Tachinidae y Erebidae, representando la dominancia de cada grupo funcional, respectivamente. Por el contrario, en la zona de menor perturbación, como fue la vegetación natural, únicamente los depredadores presentaron la mayor diversidad de insectos, con la familia Asilidae como dominante.
Agradecimientos
Los autores agradecen al Tecnológico Nacional de México por el financiamiento del proyecto “Huertos familiares y conservación de la diversidad de entomofauna benéfica” (clave: 20070.24-P) y al Consejo Nacional de Humanidades, Ciencia y Tecnología por la beca de posgrado otorgada al primer autor. También agradecemos a Xiomara Gálvez Aguilera, directora de la asociación Caribbean Conservation Coastal Ecosistem A.C., por permitirnos el acceso a la reserva privada “Komchen de los pájaros”.
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Citizen science suggests decreased diversity of insects in Mexico, a megadiverse country
Datos de ciencia ciudadana sugieren un decremento en la diversidad de insectos de México, país megadiverso
Jorge Soberon *
University of Kansas, Biodiversity Institute, Dyche Hall, 1345 Jayhawk Blvd, Lawrence, KS 66045, USA
*Corresponding author: jsoberon@ku.edu (J. Soberon)
Received: 09 July 2025; accepted: 26 August 2025
Abstract
An analysis of iNaturalist data on several taxonomic groups of insects in Mexico is presented. A decreasing trend was observed in species diversity per year for 4 families of butterflies, bumblebees, and dragonflies and damselflies. Analyses were performed on several potential vegetation types (sensu Rzedowsky), and the roles of deforestation and pesticide use on the identified trends were explored. Challenges in using unsystematic data to estimate trends are discussed, and several hypotheses are provided to explain the results.
Keywords: Butterflies; Bumblebees; Damselflies; Dragonflies, iNaturalist
Resumen
Se presenta un análisis de datos de iNaturalist sobre varios grupos taxonómicos de insectos en México. Se observó una tendencia decreciente en la diversidad de especies por año para 4 familias de mariposas, y para abejorros y libélulas. Se realizaron análisis sobre varios tipos potenciales de vegetación (sensu Rzedowsky) y se exploró el papel de la deforestación y el uso de pesticidas en las tendencias identificadas. Se discuten los desafíos del uso de datos no sistemáticos para estimar tendencias y se presentan varias hipótesis para explicar los resultados.
Palabras clave: Mariposas, Abejorros; Caballitos del diablo; Libélulas; NaturaLista
Introduction
Evidence indicates a decrease in insect populations in many countries (Edwards et al., 2025; Hallmann et al., 2017). This finding is worrisome for many reasons, including that insects are key components of ecosystems and provide societies with important ecosystem services, such as pollination (Potts et al., 2010). Moreover, insects have substantial but largely unappreciated cultural importance (Duffus et al., 2021), not only worldwide but particularly in countries such as Mexico, where insects have culinary uses (Ramos-Elorduy & Viejo-Montesinos, 2007), have been important for ancestral cultures (Beutelspacher, 1989), and have economic and societal value (Ayala et al., 2012; Rogel-Fajardo et al., 2011).
Most detailed evidence of the decline in insects has come from countries in temperate zones that have developed formal monitoring schemes (Streitberger et al., 2024; Thomas, 2005). In contrast, tropical regions are less well studied (Sánchez-Herrera et al., 2024), and the existing evidence is contradictory (Bonadies et al., 2024; Boyle et al., 2025; Wagner et al., 2021). For instance, studies on Hemiptera (Lucas et al., 2016) and on saturniid moths (Basset et al., 2017) have indicated no trends in Barro Colorado Island, Panama. Similarly, in Veracruz, Mexico, well monitored fruit flies have shown no trends (Aluja et al., 2012; Ordano et al., 2013). In contrast, decreases in saturniid larvae have been reported in Costa Rica (Salcido et al., 2020), and declines in arthropod biomass have been reported in Puerto Rico and, on the basis of a few data points, in Chamela, Mexico (Lister & García, 2018). The monarch butterfly, perhaps the best monitored insect species in Mexico, has shown consistent decreases in its wintering aggregations (Thogmartin et al., 2017; Vidal & Rendón-Salinas, 2014; Zylstra et al., 2021).
Because of its history, climate, topography, and cultural milieu (Ramamoorthy et al., 1993), Mexico is among the world’s megadiverse countries (Mittermeier et al., 1997). Therefore, assessing the trends in insect populations in Mexico should be prioritized. Unfortunately, long-term insect monitoring in Mexico is rare. Although Mexico has a long history of entomological research, including many collections and hundreds of publications (Michán & Llorente, 2002), monitoring has been limited to only a few species. Although the reasons for the lack of national monitoring schemes like those existing in other countries should be determined, this study does not attempt to do so. It takes as a premise that, in Mexico, just a few systematically obtained insect time series of more than 2-3 years long are available. This study is aimed at estimating insect biodiversity trends in Mexico, despite the absence of systematic monitoring efforts.
Systematic monitoring results are compiled in several worldwide time series databases, such as the Living Planet Index (Almond et al., 2020) and the Global Population Dynamics database (NERC Centre for Population Biology, 1999). Unfortunately, these databases have sparse insect information and contain no data for Mexico. Another possibility is using so-called citizen science (CS) data (Cohn, 2008), which, although opportunistic and unsystematic, is often abundant. Data on insects collected by non-professionals have been used to estimate phenology and distributions (Soroye et al., 2018). However, using such data to estimate population trends is challenging, as discussed below.
In Mexico, perhaps the most comprehensive CS initiative is iNaturalist (known as Naturalista in Mexico). iNaturalist began its operations in Mexico in 2008, although in 2013 the initiative came under the leadership of the national biodiversity agency, Conabio, under the name of Naturalista (Macías & Freire, 2017) and obtained funding from the Slim Foundation. Therefore, in Mexico, iNaturalist began in earnest in 2013. Despite this relatively late start, Mexico is the third country in amount of data (Mason et al., 2025) and it contains more than 80,000 records tagged as “research level” for 3 families of butterflies, and the damselflies, dragonflies, and bumblebees. This substantial information may be used to assess trends. However, CS data must be corrected for biases, of which are many (Crall et al., 2011). Specifically, in Mexico, the number of observers (and thus of observations) in iNaturalist increases each year (Table S1), and this bias should be considered when using such data.
Indeed, a major problem in using opportunistic CS data to estimate trends is correcting for biases in recording efforts (Di Cecco et al., 2021). Several methods can be used to address this problem (Isaac et al., 2014; Outhwaite, 2019; Tang et al., 2021). One of the simplest methods is correcting bias by obtaining the quotient of the metric used to report biodiversity to some measure of the effort invested in a locality, for a given period. What is “effort,” and how can it be measured in iNaturalist data? Collection effort is difficult to define but can be described in terms of: 1) the time spent collecting, 2) the method of collection and number of collectors, or 3) the number of specimens or species observed (Gulland, 1969; Willott, 2001). iNaturalist data allows for extraction of a measure of time (number of monthly observations in a year), but data quality (beyond the “research” tag, which refers to the reliability of the name assigned to the species), remains unreported, and worse, in the case of iNaturalist, this quality is known to change (Di Cecco et al., 2021). Di Cecco (2021) has suggested that, in iNaturalist, observers with at least 2 observations are more reliable than those with just 1 observation. Therefore, as a measure of effort, this study used the number of observers with 2 or more observations. As biodiversity measures it is used the number of species, and the number of observations, pooled for spatial units and year. Two indices are then calculated: number of species/effort, and number of records/effort.
Ordinary regressions of metrics against time often experience problems of autocorrelated errors and non-equal variances (heteroskedasticity). These are characteristic of time series (Shumway & Stoffer, 2005) and must be accounted for. One method of addressing the complexities of analysis of count time series data is using a package such as “trim” (in the R platform), which assumes a Poisson model for the underlying data (Pannekoek, 1998). This approach corrects for the autocorrelation of errors and for heteroskedasticity. Trim has frequently been used for European (van Strien et al., 2019) and tropical American (Novoyny & Basset, 2000) data, but the key assumption of count data (a discrete scale) complicates analysis of continuous-scale indices, or data including many non-occurrences, because the software is sensitive to the presence of zeroes, or NAs, in the data.
Another possibility is estimating whether a significant trend exists in the data, by using a non-parametric Mann-Kendall test (Lyubchich et al., 2013). An ordinary least squares linear regression (OLS) of metric against time is first performed, and the existence of trends (linear or monotonic) is subsequently determined. This method uses the sign of the slope in the OLS, and the significance is tested with the Mann-Kendall test.
Additional methods can be used, such as, for a single species, the logit of the probability of occupancy of a cell, on a time unit (van Strien et al., 2019) and fit a generalized linear model of predictors, by using the length of the list of species as a measure of effort (Szabo et al., 2010). Then several single species regressions can be combined in an index (van Strien et al., 2019). One problem with this approach is that generalized linear modeling is based on an assumption of independence of errors, which might be violated in a time series.
A statistically more sophisticated modification of the above idea is reporting the proportion of occupied sites under a hierarchical model that separates the actual presence from the act of observation (Outhwaite, 2019). Although apparently very rigorous, this approach has its own problems, including the need to define an appropriate model for the “present” and “observer” components, and the need to have replicated visits to the same site within the same season (van Strien et al., 2019).
Another possibility is using generalized least squares (GLS) regressions, which allow for autocorrelated errors and heteroskedasticity. The R package “nlme” implements this technique. This method can fit an ordinary regression of the index against covariates such as time, and another regression including autocorrelation with power variance decay in its model. Subsequently, the 2 models can be compared with the Akaike criterion, and the best model can be retained. This option was used here, with the simplest ARIMA model with lag = 1 as a model of correlated errors.
This study further assessed whether any existing trends might have differed for different ecological regions of Mexico. A variety of subdivisions of Mexico have been suggested, according to different ecological perspectives, at different spatial resolutions (Anonymous, 1997; Challenger & Soberón, 2012; Miranda & Hernández, 1963; Olson et al., 2001). Here, Rzedowsky’s potential vegetation types were used (Rzedowsky, 1986). Although coarse-grained, these types are based primarily on straightforward floristic criteria, are well known in Mexico, and have a small number of categories.
An important caveat in using CS data is that species that are difficult to identify by sight should be avoided. This work focused on 3 families of butterflies (Papilionidae, Pieridae, and Nymphalidae), with 316 names (skippers and the smallest families in the Papilionoidea were excluded); 23 names for bumblebees; and 293 names for the Odonata (both Zygoptera and Anysoptera). In addition, as a comparison, data on 307 names for Solanaceae were included. The numbers of names (without proper taxonomic validation by experts), as reported by iNaturalist, are listed in Table 1.
For the butterflies, although using species identified as indicators of “conservation” status (Orta et al., 2022) would have been interesting, most species identified by these authors as indicators had only a few records in the iNaturalist database. Therefore, the analysis was performed not by individual species, but by pooling all the data in the 3 families of butterflies, all the dragon and damselflies, and all the bumblebees.
For obtaining uncertainty bands, grids of hexagons covering the territory of Mexico were defined at several resolutions (Fig. 1). For a given year and taxonomic group, the means and variances over hexagons were determined. Each unique combination of year and hexagon defined an “event,” and thus the abundance metrics were: 1) the number of observations per event (cumulative monthly observations); and 2) the number of different species per event. As a measure of effort, the total number of different observers with at least 2 observations in each “event” was used. The final index was the average over all hexagons with at least 1 record, for a given year, of the number of observations or the number of species, per observer.
Table 1
Numbers of scientific names for the different taxonomic groups in the 4 most visited potential vegetation classes. The butterflies are the Papilionidae (swallowtails), Pieridae (sulfurs), and Nymphalidae (brushfoots).
| All Mexico | Xerophytic Shrub | Pine Oak Forest | Grasslands | Tropical Deciduous Forest | |
| Butterflies | 413 | 153 | 220 | 220 | 198 |
| Odonata | 292 | 185 | 187 | 105 | 184 |
| Bombus | 23 | 16 | 20 | 9 | 14 |
| Solanaceae | 307 | 176 | 210 | 111 | 153 |
Changing the hexagon area might potentially change the results. This problem, described as the “modifiable areal unit problem,” has been long known to geographers (Openshaw, 1984). Fortunately, in this case, the qualitative results were not affected by the resolution of the hexagons (data correlations among resolutions always exceeded 0.7). Consequently, only the analysis using the largest (2 degrees) hexagons (n = 81) is reported.
The literature has suggested that the decrease in insect abundance has been due to: 1) increased use of pesticides, 2) decreased habitat area (or increased transformed land area), and 3) climate change. At the scale of the whole country, regressions of data versus time series of pesticide use and deforestation rates are reported.
Materials and methods
CS data are not ideally suited to the estimation of trends, primarily because of the biased and uneven methods of sampling sites, times, and species. This work used 1 of the 3 methods proposed by Isaac et al. (2014): correcting the reported number of sightings according to a measure of effort. iNaturalist data were downloaded from the Global Biodiversity Information Network (GBIF), as detailed in Table 2.
Data were divided into subsets (keeping records with coordinates) for the 4 largest families in the Papilionoidea: Papilionidae (5,583 records), Pieridae (29,994 records), Nymphalidae (20,145 records), and Lycaenidae (2,326 records). The Lycaenidae was removed from the analysis because many species are relatively difficult to determine visually. Data for the genus Bombus (bumblebees, 6,543 records) and the 2 suborders of the Odonata (the Zygoptera, 12,772 records, and the Anisoptera, 24,003 records) were also downloaded. For comparison purposes, observations of the nightshade family, the Solanaceae (39,014 records), were downloaded. The number and positions of every observation in Mexico are presented in figure 2.
Table 2
Digital Object Identifiers (DOIs) from GBIF for the datasets used in the work.
| Taxon | GBIF DOI | iNaturalist Records | Unique names |
| Nymphalidae | doi.org/10.15468/dl.qta4zp | 20,145 | 223 |
| Papilionidae | doi.org/10.15468/dl.uu4unc | 5,583 | 13 |
| Pieridae | doi.org/10.15468/dl.zug4ee | 29,994 | 80 |
| Lycaenidae | doi.org/10.15468/dl.xgad6g | 2,326 | 97 |
| Odonata | doi.org/10.15468/dl.atdkfz | 36,775 | 292 |
| Anisoptera | 24,003 | 163 | |
| Zygoptera | 12,772 | 129 | |
| Bombus | doi.org/10.15468/dl.c6h4jz | 6,543 | 23 |
| Solanaceae | doi.org/10.15468/dl.597nj5 | 39,014 | 307 |
Data tagged as “research quality” in the downloaded GBIF data were retained, and basic data cleaning was performed to keep the coordinates inside Mexico. No attempt was made to correct for outdated taxonomy or other known issues present in aggregator data (Chapman, 2005).
Data can be organized as time series, by pooling the observations in a year. This method has a drawback of potentially missing seasonality; however, pooling by month produces tables that are too sparse and therefore are difficult to analyze. To include some measure of uncertainty in the trends, the averages of the calculated indices over all non-empty (i.e., with at least 1 observation) hexagons of 2 degrees of surface were determined, and its standard error calculated.
Two indices were used: different_species/observer and records/observer. The first is a measure of diversity, whereas the second is a measure of abundance. Findings for both are reported. “Observers” refers to the number of observers with at least 2 registered observations.
To summarize trends, a useful statistic may be the slope of a linear model of index as a function of time, which requires regressions of index vs. year. However, as previously discussed, the errors in many time series are not independent, and the equal variance assumption of ordinary least squares is also often violated. If uncorrec-
ted, these problems interfere with rigorous calculations of probability under a null hypothesis (McShane et al., 2019). Among the many methods for addressing this problem, generalized least squares regressions (Baillie & Kim, 2018), which enable inclusion of an autoregressive structure of correlations and violations of homoscedasticity, were chosen herein. Two models were fitted to the data: an ordinary linear least squares, and a first order auto-regressive, moving average model (ARIMA) (Shumway & Stoffer, 2005) allowing for heteroskedasticity. The 2 models were compared with an ANOVA (Fox & Weisberg, 2019), and the most likely model (based on the Akaike criterion; see Fox & Weisberg, 2019) was used. This process permitted to obtain, in a rigorous way, the probability for the observed slope values, under a null hypothesis of a slope equal to zero. Reporting the “significance” of slopes has been substantially criticized (McShane et al., 2019). Therefore, the probability (rather than the “significance”) of the slope, based on the assumption of a null model of no trend, is reported. Very small probabilities are highlighted.
The regressions included 2 possible causal factors: forest loss and use of pesticides. The deforestation rate was obtained from the Global Forest Watch website (Sims et al., 2024) with a threshold of 30% of forest cover, as recommended by Sims et al. (2024). This dataset has maintained methodological consistency (Hansen et al., 2013) and therefore is preferable to the INEGI Series (Gebhardt et al., 2015). Agrochemical use was determined as the amount of pesticides used per hectare of cropland, as reported on the FAO Web site. The data came from government reports https://www.fao.org/faostat/en/#data/RP. A discussion of the FAO dataset’s strengths and problems has been provided by Shattuck et al. (2023).
Because the probability of the observed values of the slope of the index of diversity per unit of effort vs. time, under a null hypothesis of 0 slope, was small in most cases, the regression was assumed to remove the time trend, and factors affecting just the residuals were searched for. That is, the residual of the index vs. time regressions was regressed against 2 predictors: deforestation rate and use of pesticides. The results are shown in the Supplementary materials.
To aggregate by “biome,” the subdivision of the Potential Vegetation of Mexico (Rzedowsky, 1986) was selected. A shapefile of Rzedowsky’s map at 1:4,000,000 scale, available at Conabio Geoportal, is produced by Instituto de Geografía, UNAM México. This map was used to pool the iNaturalist records according to potential vegetation, by using the 4 categories with the highest number of iNaturalist reports.
An informal survey was circulated among scientists working in 3 major ecology research centers in Mexico (INECOL, Veracruz, Instituto de Ecología, UNAM, and Ecosur, Chiapas). A total of 37 questionnaires were sent with Qualtrics. The questions are provided in the Supplementary materials. The main data tables and R code are openly available (Creative Commons CC0: 1) at https://github.com/jsoberon/iNaturalistInsectsMexico
Results
The informal questionnaire received 27 responses out of 37 requests. Among the respondents, 84% stated that they have observed a decrease in the number of insects either in streetlights in villages, or in the windshields or radiators of field vehicles. Although these answers lacked statistical rigor, they suggested a widespread perception among field biologists in Mexico that insect populations are becoming smaller.
The iNaturalist data provided a more nuanced picture. Before examining the trends in biodiversity indices, basic data were analyzed. Indeed, both the number of species and the number of observers (with more than 2 observations) increased (Fig. 3).
The numbers of observed species and observers both increase over time. The increased number of observers introduced an important bias in the data, given that more species (or more individuals) would reasonably be expected to be reported if more observers were present. However, although the diversity of insects appeared to be decreasing, the evidence of a decrease in abundance was unclear (Fig. 4; Tables 3, 4).
Diversity per unit effort appeared to decrease (Table 3). However, the trends in the abundance (observations/number observers) were either positive or indistinguishable from 0 (Table 4). Box plots of the slopes of the regressions for the 2 indices (species, and observations) are shown in figure 5.
The above results suggest that diversity is decreasing, but abundance is stable. This finding is inconsistent with the informal perceptions of field biologists (as indicated by the questionnaire), most of whom perceived diminished insect abundance. Among the few insect species whose abundance in Mexico has been monitored systematically, Danaus plexippus (monarch butterfly) populations are decreasing (Vidal & Rendón-Salinas, 2014; Zylstra et al., 2021), whereas Anastrepha fruit fly populations appear to be stable (Aluja et al., 2012; Ordano et al., 2013). Comparing these 2 cases is challenging, because monarch butterflies are affected by a variety of factors occurring on a continental scale, whereas fruit flies might be affected primarily by local factors.
Might the negative trend in diversity correlate with predictors often associated with insect loss? Forest cover, as measured via remote sensing over 15 years (Hansen et al., 2013), is decreasing in Mexico (Supplementary materials). Pesticide use per hectare of crop, as reported by the FAO, increased until 2018, when the FAO database indicated an abrupt decrease (Supplementary materials). The causes of this decrease, if real, are unknown; however, after the COVID-19 pandemic, Mexico’s primary sector experienced a marked decrease in activity (Sánchez et al., 2022), which may explain a drop in the use of agrochemicals. Regressions of the residuals of the diversity/effort vs. time models against 2 predictors, deforestation rate and pesticide use per hectare, were not associated with small probabilities of an H0 of 0 slope (Supplementary materials). Consequently, the data did not provide evidence that negative slopes in insect diversity were due to pesticide use or deforestation.
Finally, for the major taxonomic groups, the slopes of the generalized least squares, in the first 4 potential vegetation types according to Rzedowsky (1986) were most negative for bumblebees in tropical deciduous forest, followed by pine-oak forest and xerophytic shrub. For the butterflies, the most negative slope was in pine-oak forest, followed by tropical deciduous forest and xerophytic shrub (Supplementary materials: Table S3). In the case of the Solanaceae, a group included for comparison purposes, the slope is only negative in the grasslands vegetation type.
Table 3
Regression analysis (generalized least squares) of diversity/observer vs. time in the iNaturalist data, for the main taxonomic groups. The analysis was performed over the mean values in hexagons of 2 degrees of resolution. With the exception of the Zygoptera, for which the first order autoregressive model did not converge, the ordinary least squares regression did not significantly differ with respect to models with autocovariance and heteroskedasticity. Consequently, the table shows the slope of ordinary linear models of different_species/effort with respect to time. The probabilities of the obtained values under a null hypothesis of slope of zero were very small, with the exception of the dragonflies and swallowtails (Fig. 2).
| Taxon | Species, 2 degrees | |||
| Slope | p | Model | n | |
| Bombus | -0.0428 | 0.0000237 | OLS | 6,543 |
| Anisoptera | -0.0206 | 0.00155 | OLS | 24,003 |
| Zygoptera | 0.0003 | 0.942 | OLS_NO_CNV | 12,772 |
| Nymphalidae | -0.0372 | 0.000000242 | OLS | 20,145 |
| Papilionidae | -0.0108 | 0.104 | OLS | 5,583 |
| Pieridae | -0.0247 | 0.000202 | OLS | 29,994 |
Table 4
Regression analysis (generalized least squares) of records/observer vs. time in the iNaturalist data, for the main taxonomic groups. The data were averaged over hexagons of 2 degrees of resolution. Except for the Zygoptera, for which a first order autoregressive model was used, the ordinary least squares regression did not significantly differ with respect to models with autocovariance and heteroskedasticity. Consequently, the table shows the slope of ordinary linear regressions of number_of_records/effort with respect to time. Notably, every regression had a positive, low probability slope.
| Taxon | Slope | p | Model | n |
| Bombus | 0.0446 | 0.00000766 | OLS | 6,543 |
| Anisoptera | 0.0201 | 0.00649 | OLS | 24,003 |
| Zygoptera | 0.0366 | 6.86E-08 | ARIMA | 12,772 |
| Nymphalidae | 0.0089 | 0.034 | OLS | 20,145 |
| Papilionidae | 0.0669 | 0.00000176 | OLS | 5,583 |
| Pieridae | 0.0154 | 0.0000984 | OLS | 29,994 |
Discussion
The results show a tendency to decrease the number of species with time, for the insects, and a much less marked negative trend for the Solanaceae. This suggests that CS data does capture some sort of biological signal in the data. However, a diminishing trend of diversity, together with a stable pattern of abundance, are compatible with several hypotheses. One entirely biological hypothesis is that insect diversity, but not abundance, is decreasing. If the rarest species are disappearing, then the country is homogenizing (McKinney & Lockwood, 1999). Thus, Mexico’s highly diverse and unique insect biodiversity is slowly being replaced by a more homogeneous, more cosmopolitan set of species. This rather alarming possibility, supported by the CS data, must be more directly assessed in the field.
Another explanation for the observed negative trend in insect species numbers might be that, over time, observers have reached the asymptote of the total number of species available to be observed. Since the total number of species in any given area is probably roughly constant, with sufficient effort, no more than that constant number can be reported; however, if the number of observers is increasing, a negative trend in the index of species/observers would result. The total number of species in the database, for each taxonomic group, is shown in Table 1. The average number of species reported per hexagon was well below that total (Supplementary materials: Table S4), thus suggesting that a saturation effect was not present, and the results presented here indeed indicate a decreasing trend in insect diversity. This complex point is discussed at more length in the Supplementary materials.
Finally, the negative trend is also compatible with a hypothesis regarding the quality of iNaturalist observers in which the number of observers has increased, as indicated by the data, whereas the observers’ discrimination ability or interests might have changed over time, perhaps because they focused on common species. Unfortunately, the very nature of the information in CS data makes assessing this effect very difficult. This aspect essentially describes the main problem with using unstructured CS data: because the methods are not standardized, any trend in the data might be explained by a trend in the behavior of the observers.
What explanations can be deduced for the absence of trends in the number of observations/effort? One possibility is that the presence of more observers simply resulted in more observations, and the number of observations and observers with more than 2 observations are roughly proportional. This means that the lack of trend could be an artifact of the data.
These results should be considered as hypotheses to be examined through more direct methods. Nonetheless, the results strongly suggest decreasing numbers of species in butterflies (important from a cultural perspective and perhaps a pollination perspective), bumblebees (important as pollinators), and Odonata (important as insect predators and as indicators). Therefore, the biodiversity of some of the most important and underappreciated groups of species in Mexico appears to be decreasing. If confirmed, this result would be highly alarming. Indeed, insects are key components of ecosystems (Noriega et al., 2018). Although for most insect species in Mexico we do not have direct documentation of their role, or of the economic and cultural value of their services, we have substantial indirect evidence of their importance as pollinators (Ashworth et al., 2009), as natural enemies of agricultural pests (López et al., 1999; Aluja et al., 2014), and as potential for non-conventional agri-business (López-Gutierrez et al., 2023). If substantiated, the decrease we report should be a major cause of alarm for Mexicans.
What might be causing a decrease? In a study in Europe (Schuch et al., 2012), in which a similar decrease in diversity was reported in a family of bugs of agricultural importance, a concomitant loss of non-agricultural habitat for the insects was reported. The study was conducted at the species level, and the authors argue that the more specialized, less tolerant species are those disappearing because of agricultural expansion. Again, monitoring using standardized procedures is required to test this idea.
Climate change is often cited as a cause of population decline in insects. However, climate change is a long-term phenomenon that occurs at the scale of many decades. To demonstrate climate change as a factor affecting population size, modeling or documentation of the effect of mean and variance in climatic variables on long-term population time series is necessary (Batalden et al., 2014; Boggs, 2016). The data used in this work is not appropriate for this purpose.
The results suggested that negative trends might not be identical among ecological regions. However, interestingly, the iNaturalist data indicated that pine-oak forest, xeric shrub, and tropical deciduous forest might be hotspots of diversity loss. This finding is somewhat surprising, given the widespread concern regarding tropical rainforests. Of course, the results may be due to the scarcity of data for tropical wet vegetation types.
CS data exists in substantial and growing amounts. It is a very valuable source of data. However, as with any data, it contains biases that are sometimes difficult to remove. The unavoidable conclusion is that Mexico must crucially invest in countrywide insect monitoring schemes based on systematic methods. Several approaches could be used. The first is improving CS schemes, providing training, and applying standard protocols, as performed in Canada, the USA, and many European countries (Streiter et al., 2024). This approach might be useful only for conspicuous, easily identifiable species, yet it markedly influences public environmental awareness and therefore should be maintained (Dickinson et al., 2012). Several methods using advanced technologies include computer vision, bioacoustics, and metagenomics can also be used (Van Klink et al., 2022). For bats, monitoring is already underway in Mexico (Zamora-Gutiérrez et al., 2020). Adoption of high technology methods would require funding, training, and substantial analytical capacity.
Regardless of the method chosen, in Mexico, the fourth most biologically diverse country on the planet, monitoring as many biodiversity components as possible is critical, and insects, “the little things that run the world” in the words of E. O. Wilson, appear to be disappearing very rapidly. Societies need to pay attention.
Acknowledgements
I am very grateful to Luis Eguiarte and Rodrigo Medellín, of the Mexican Instituto de Ecología, for their detailed and positive criticism of the methods I used. Their feedback led to several changes in the data analysis methods. Exequiel Ezcurra, of the University of California at Irvine, and Carlos Martinez, formerly of the University of Wyoming, also made very helpful statistical suggestions. My students Jennifer Ramos and Anahí Quezada helped me download and organize the data and patiently discussed the project with me. I gratefully thank the field ecologists in Mexico who took the time to respond to the questionnaire regarding their experiences with insects in the field.
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Claves ilustradas para la identificación de especies y tribus de Scolytinae (Coleoptera: Curculionidae) de la provincia de Tucumán, Argentina
Illustrated keys for the identification of species and tribes of Scolytinae (Coleoptera: Curculionidae) from Tucumán Province, Argentina
Silvia Patricia Córdoba a, * y Thomas H. Atkinson b
a Fundación Miguel Lillo, Instituto de Entomología, Área de Zoología, Miguel Lillo 251, 4000 San Miguel de Tucumán, Argentina
b University of Texas at Austin, University Texas Insect Collection, Lake Austin Center 3001 Lake Austin Boulevard, Suite 1.314, Austin, 78703 Texas, EUA
*Autor para correspondencia: spcordoba@lillo.org.ar (S.P. Córdoba)
Recibido: 08 abril 2025; aceptado: 25 septiembre 2025
Resumen
La subfamilia Scolytinae es reconocida por comprender especies de importancia forestal, frutícola y ornamental debido a su impacto económico y por poseer la función, desde el punto de vista ecológico, de regular las poblaciones vegetales con las que se asocian. Para la Argentina, no existen claves para la determinación de las tribus y especies, por lo que en el presente trabajo se incluye una para la identificación de las tribus y 12 claves para la identificación de 56 especies, como así también su distribución en América del Sur y Argentina. También se añaden fotografías del aspecto general y caracteres más importantes de cada especie.
Palabras clave: Escarabajos descortezadores; Escarabajos de ambrosía; América del Sur; Tucumán; Distribución; Forestal
Abstract
The subfamily Scolytinae is recognized for comprising species of forestal, fruit, and ornamental importance due to their economic impact and their ecological role in regulating the plant populations with which they are associated. For Argentina, there are no keys for identifying the tribes and species; therefore, this work includes a key for identifying the tribes and 12 keys for identifying 56 species, as well as their distribution in South America and Argentina. Photographs of the general appearance and most important characteristics of each species are also included.
Keywords: Bark beetles; Ambrosia beetles; South America; Tucumán; Distribution; Forestry
Introducción
La subfamilia Scolytinae constituye un grupo grande y diverso de escarabajos barrenadores (Atkinson, 2017). Presentan una amplia gama de modos de alimentación, pero se dividen principalmente en 2 grupos: los escarabajos descortezadores y los de ambrosía. Los primeros se alimentan del floema de árboles enfermos o moribundos y por su actividad, la corteza termina por desprenderse. Los escarabajos de ambrosía perforan el xilema de árboles debilitados, moribundos o enfermos y cultivan hongos simbiontes, los cuales constituyen su alimento (Kirkendall et al., 2015). La subfamilia es reconocida por comprender especies de importancia económica y por poseer la función, desde el punto de vista ecológico, de regular las poblaciones vegetales con las que se asocian (Pérez-De La Cruz et al., 2016). Además, los escarabajos descortezadores son importantes, ya que fomentan la sucesión, lo que conlleva a mantener la salud forestal en muchos ecosistemas del mundo (Morris et al., 2018). Aunque la mayoría se encuentra en árboles moribundos, debilitados, enfermos o trocería recién cortada, algunas especies pueden constituir verdaderas plagas al invadir árboles sanos o sin evidencia de debilitamiento (Lombardero, 1995; Pérez-Silva et al., 2021). Su peligrosidad radica, además, en que algunas especies son vectores de hongos fitopatógenos, los cuales pueden ocasionar la muerte de la planta o producir daños que alteren la calidad de la madera provocando pérdidas económicas importantes (Lombardero, 1995). Además, por la naturaleza críptica de las galerías, son difíciles de detectar y esto conlleva a que muchas especies se hayan establecido en otras áreas fuera de su rango de distribución, estableciéndose como especies invasoras, planteando graves amenazas a los bosques, productos forestales y cultivos (Kirkendall, 2018).
Hasta el año 2018 en Argentina y, particularmente en la provincia de Tucumán, gran parte de los estudios sobre Scolytinae corresponden a trabajos dispersos en la literatura (Bruch, 1914; Bosq, 1943; Hayward, 1960; Santoro, 1966; Viana, 1964; Wood, 2007). Hasta la fecha se han descrito unas 6,400 especies en todo el mundo (Hulcr et al., 2015; Wood y Bright, 1992). Para Argentina, Atkinson (2025) lista 17 tribus, 52 géneros y 199 especies, de las cuales 18 son introducidas en el país desde otros continentes. Para la provincia de Tucumán, se han reportado 56 especies distribuidas en 12 tribus y 26 géneros (Córdoba et al., 2025).
Hasta el momento, no se han desarrollado claves para la identificación de tribus y especies en Argentina, lo cual limita la capacidad de diagnóstico y manejo de estos insectos. El presente trabajo constituye la primera iniciativa dirigida a la elaboración de claves ilustradas con figuras, que permitan la determinación precisa de las tribus y especies presentes en el país. Estas herramientas son esenciales, ya que algunas especies tienen un impacto significativo en cultivos forestales, ornamentales y frutales, que afectan tanto la economía nacional como la biodiversidad regional. En un contexto de cambio climático y globalización, la identificación temprana y precisa de estas especies resulta esencial para prevenir posibles brotes. La ausencia de herramientas específicas adaptadas a las condiciones locales hace que este trabajo sea aún más relevante al proporcionar un recurso indispensable para investigadores, técnicos y productores.
Materiales y métodos
Gran parte del material estudiado fue obtenido de recolectas directas e indirectas, realizadas desde el 2016 al 2024, en la provincia de Tucumán, noroeste de Argentina. Las recolectas indirectas se realizaron utilizando trampas elaboradas con botellas plásticas y con etanol 96% como atrayente. Los ejemplares fueron separados y determinados siguiendo las claves de Wood (2007), así como por comparación con los ejemplares de distintas colecciones y la revisión por parte del segundo autor de los especímenes del Smithsonian National Museum of Natural History (USNM), Washington, DC (EE.UU.) y el Naturhistorisches Museum Wien (NHMW), Viena, Austria.
La elaboración de las claves se realizó con base en las descripciones de Wood (2007) y en la revisión de los ejemplares de las colecciones de la Fundación Miguel Lillo (I-FML), Museo de La Plata (MLP) y Museo Argentino de Ciencias Naturales Bernardino Rivadavia (MACN). En las claves fueron incluidos mayormente caracteres de las hembras, sin embargo, en algunos casos se analizaron los caracteres del macho, por lo que se aclara con (Macho), al inicio de cada llave. Las especies marcadas con un asterisco (*) son especies exóticas. La distribución de las especies en el país se realizó con base en la información obtenida de las colecciones revisadas, la literatura existente, datos de las recolectas y en la información presente en el sitio web Bark and Ambrosia Beetles of the Americas (Atkinson, 2025). Las claves incluyen especies que se distribuyen en otras provincias de Argentina, por lo que pueden ser utilizadas en otras regiones del país.
Las fotografías fueron tomadas con una cámara Leica DMC 2900 incorporada a un microscopio estereoscópico Leica M205 C, para cada foto se tomó una serie de capas a través del Software Leica Application Suite Core versión 4.7.1.
Resultados
Se identificaron 56 especies pertenecientes a 25 géneros y 12 tribus: Bothrosternini, Corthylini, Dryocoetini, Hexacolini, Hylurgini, Ipini, Micracidini, Phloeosinini, Phloeotribini, Scolytini, Trypophloeini y Xyleborini. La tribu más diversa fue Xileborini, con 14 especies, seguida de Bothrosternini, con 12 especies.
Los géneros históricamente incluidos dentro de Cryphalini fueron agrupados de acuerdo con la propuesta de Johnson et al. (2020), Hypothenemus Westwood, 1836 se incluye dentro de la tribu Trypophloeini, y Acorthylus y Cryptocarenus se incluyen dentro de la tribu Corthylini (subtribu Pityophthorina). Además, se incluyen tanto a Xyleborus bispinatus Eichhoff, 1868 como a X. ferrugineus (Fabricius, 1801) como 2 especies distintas de acuerdo con Kirkendall y Jordal (2006). Theoborus theobromae (Hopkins,1915) se trata como Coptoborus villosulus (Blandford, 1898), de acuerdo con Smith y Cognato (2021).
La clave para tribus es una clave simplificada que no se puede utilizar fuera del área de estudio, en gran parte porque omite grupos no presentes en el noroeste de Argentina. Las claves para tribus y géneros de Wood (1982, 2007) no permiten ubicar todos los géneros en tribus con certeza. En particular, los géneros Acorthylus y Cryptocarenus fueron tratados por Wood (1982, 2007) en su tribu Cryphalini, pero autores más recientes como Johnson et al. (2020) los han incluido en la tribu Corthylini (subtribu Pityophthorina). En esta clave se identifican por separado.
Clave de las tribus de Scolytinae presentes en Tucumán
1. Cabeza parcialmente cubierta por el pronoto; pronoto en vista lateral con curvatura gradual desde la base hasta el ápice (fig. 1a-d) 2
1’. Cabeza completamente cubierta por el borde anterior del pronoto (mejor visto en aspecto lateral); pronoto en vista lateral generalmente con un cambio abrupto en armadura o contextura (fig. 1e- h) 7
2. Protibia sin dentículos laterales, con gancho apical en borde exterior (fig. 2a) Scolytini
2’. Protibia con dentículos laterales, proyecciones apicales pueden estar presentes en borde interior (fig. 2b-d) 3
3. Protibia con 1 proyección subapical en borde exterior, consistiendo en 2 dentículos prominentes (fig. 2b) Bothrosternini
3’. Protibia sin proyección subapical en borde exterior (fig. 2c, d) 4
4. Maza antenal con segmentos, independientemente móviles (fig. 1c, d) Phloeotribini
4’. Maza antenal con segmentos fusionados o sin segmentación (fig. 2f-n) 5
5. Funículo se une a la maza antenal en el centro de la base del mismo (fig. 2f) Hylurgini
5’. Funículo se une a la maza antenal lateralmente (fig. 2g, h) Phloeosinini
6. Perfil lateral del pronoto curvado, cima poco desarrollada (fig. 1e, f) 7
6’. Perfil lateral del pronoto con cima pronunciada y abrupta (fig. 1g, h) 8
7. Protibia con 2 dentículos prominentes ubicados en el ápice (fig. 2e); funículo antenal con 7 segmentos Hexacolini
7’. Protibia con dentículos abundantes, similares en tamaño (fig. 2a, b), funículo antenal con 5 segmentos Dryocoetini
8. Parte terminal del dorso del último segmento abdominal parcialmente visible en vista ventral (abdomen parece tener 6 segmentos en vista ventral) (fig. 1j) Corthylini (Corthylini, Pityophthorini en parte)
8’. Parte terminal del dorso del último segmento abdominal parcialmente no visible en vista ventral (abdomen parece tener 5 segmentos en vista ventral) (fig. 1i) 9
9. Zona posterior de la cavidad oral deprimida debajo del nivel de la cabeza en vista ventral (fig. 1k) Xyleborini
9’. Zona posterior de la cavidad oral no deprimidas (fig. 1l) 10
10. Suturas de maza antenal similares en caras anterior y posterior (fig. 2k, l) 11
10’. Suturas de maza antenal conspicuamente desplazadas en cara posterior (oblicuamente truncada) (fig. 2m, n) 13
11. Segundo segmento del flagelo antenal tan largo como segmentos 3 y 5 juntos (fig. 2i) Acorthylus (Corthylini: Pityopthorina
11’. Segundo segmento del flagelo antenal igual de largo como segmentos 3 y 5 12
12’. Vestidura de las interestrías del disco elitral escasa o ausente, setas en forma de cuchara (fig. 2o) Cryptocarenus (Corthylini: Pityopthorina)
12’. Vestidura de las interestrías abundante desde la base del disco hasta el ápice del declive; setas en forma de cinta, truncadas en su ápice (fig. 2p) Trypophloeini
13. Funículo antenal con 6 segmentos (fig. 2j) Micracidini
13’. Funículo antenal con 5 segmentos Ipini
Tribu Bothrosternini Blandford, 1896
Presentan dimorfismo sexual en la frente, los ojos son enteros, el funículo antenal lleva 6 segmentos, la maza antenal es moderadamente aplanada y con suturas bien marcadas, las procoxas están moderadamente o ampliamente separadas, el pronoto carece de armaduras o lleva pocas ornamentaciones, las protibias presentan un proceso bífido en el ángulo apical externo que excede el ángulo apical interno, la base de los márgenes de los élitros generalmente llevan una fila de crenulaciones poco desarrolladas, representada en algunas especies, por una costa continua. Son monógamos, con excepción del género Bothrosternus, donde ocurre algún tipo de partenogénesis. La mayoría de las especies son barrenadoras de la médula de ramas pequeñas, con excepción de Pagiocerus que barrena semillas y Bothrosternus que también barrena ramitas, pero aparentemente es xylomicetófago. Esta tribu se restringe a zonas tropicales de América, solo una especie se extiende hasta el sureste de EUA (Wood, 2007).
Clave de las especies de la tribu Bothrosternini
1. Cuerpo alargado y delgado (fig. 3a, c); vestidura que incluye setas o escamas; suturas de la maza antenal rectas y transversales (fig. 3e); ancho del rostro mayor que distancia entre los ojos; pronoto longitudinalmente estriado y con puntos; principalmente barrenadores de ramas y tallos delgados 2
1’. Cuerpo ovalado y robusto (fig. 3b, d); vestidura compuesta de setas; suturas de maza antenal fuertemente procurvadas (fig. 3f); ancho del ápice del rostro igual a distancia entre los ojos; pronoto con puntuación bien marcada; barrenadores de semillas Pagiocerus frontalis (Fabricius)
2. Presencia de crenulaciones en las zonas anterolaterales del pronoto; pronoto casi tan ancho como largo; ancho del ápice del epistoma mayor que distancia entre los ojos 3
2’. Superficie de pronoto lisa; pronoto más largo que ancho (fig. 4b); ancho del ápice del epistoma igual a distancia entre ojos Cnesinus dividuus Schedl
3. Vestidura compuesta por escamas y setas grisáceas (fig. 4a, c); lados del pronoto paralelos (fig. 5a); frente con abundante pubescencia, compuesta de setas blanquecinas, apretadas contra el fondo y dirigidas hacia el centro (fig. 4e) Cnesinus squamifer Wood
3’. Vestidura compuesta de setas amarillentas (fig. 4b, d); lados del pronoto convexos (fig. 4b); frente con escasa pubescencia, setas amarillentas, semierectas y dirigidas hacia el centro (fig. 4f) Cnesinus hispidus Eggers
Pagiocerus frontalis (Fabricius, 1801) (fig. 3b, d, f)
Distribución. Argentina: S/D, Tucumán; América del Sur: Bolivia, Brasil, Chile, Colombia, Ecuador, Guyana Francesa, Paraguay, Perú, Trinidad y Tobago y Venezuela (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Smith et al., 2017; Wood, 2007).
Comentarios. Costilla y Coronel (1994) citan a esta especie como Pagiocerus fiorii Eggers sin especificar ninguna ubicación. Está extensamente distribuida en regiones tropicales de América.
Cnesinus dividuus Schedl, 1938 (fig. 3a, c, e)
Distribución. Argentina: Buenos Aires, Tucumán; América del Sur: Brasil (Atkinson, 2025; Córdoba et al., 2025; Wood, 2007).
Cnesinus squamifer Wood, 2007 (fig. 4a, c, e)
Distribución. Argentina: Tucumán (Atkinson, 2025; Córdoba et al., 2023; Wood, 2007).
Cnesinus hispidus Eggers, 1943 (fig. 4b, d, f)
Distribución. Argentina: Tucumán; América del Sur: Bolivia y Brasil (Atkinson, 2025; Córdoba et al., 2021, 2023, 2025; Wood, 2007).
Tribu Corthylini LeConte, 1876
Por lo general machos y hembras similares en tamaño. Presentan un surco en el extremo anterior del metapisternon que conforma un mecanismo de bloqueo que mantiene los élitros en su lugar cuando están en reposo. La maza antenal es aplanada, con las suturas similares en ambos lados de la misma y los de la cara posterior no están fuertemente desplazados hacia el ápice. Las tibias son delgadas y llevan pocos dentículos en el margen lateral. Esta tribu se divide en 2 subtribus: Corthylina y Pityophthorina (Wood, 2007).
Figura 1. a, Vista lateral de cabeza y pronoto de Scolytopsis toba;b, vista dorsal de cabeza y pronoto de S. toba; c, vista lateral de cabeza y pronoto de Phloeotribus harringtoni; d, vista dorsal de cabeza y pronoto de P. harringtoni; e, vista lateral de cabeza y pronoto de Scolytodes tucumani; f, vista dorsal de cabeza y pronoto de S. tucumani; g, vista lateral de cabeza y pronoto de Orthotomicus laricis; h, vista dorsal de cabeza y pronoto de O. laricis; i, vista ventral del abdomen de Xyleborus volvulus; j, vista ventral del abdomen de Araptus sp.; k, vista ventral de partes bucales de Xylosandrus crassiusculus; l, vista ventral de partes bucales de Araptus pubescens.
Figura 2. a, Protibia de Scolytopsis puncticollis; b, protibia de Pagiocerus frontalis; c, protibia de Coccotrypes carpophagus; d, protibia de Orthotomicus laricis; e, protibia de Scolytodes tucumani; f, antena de Xylechinus imperialis; g, antena de Chramesus argentiniae; h, antena de Pseudochramesus acuteclavatus; i, antena de Acorthylus bosqui; j, antena de Hylocurus giganteus; k, vista de cara anterior de la antena de Gnathotrichus sulcatus; l, vista de cara posterior de la antena de G. sulcatus; m, vista de cara anterior de la antena de Xyleborus volvulus; n, vista de cara posterior de la antena de X. volvulus; o, disco y declive elitral de Cryptocarenus seriatus; p: disco y declive elitral de Hypothenemus crudiae.
Figura 3. a: Vista dorsal de Cnesinus dividuus (tomada de Atkinson, 2025); b, vista dorsal de Pagiocerus frontalis (tomada de Atkinson, 2025); c, vista lateral de C. dividuus (tomada de Atkinson, 2025); d, vista lateral de P. frontalis (tomada de Atkinson, 2025); e, antena de C. dividuus;f, antena de P. frontalis.
Figura 4. a, Vista dorsal de Cnesinus squamifer (tomada de Atkinson, 2025); b, vista dorsal de C. hispidus (tomada de Atkinson, 2025); c, vista lateral de C. squamifer (tomada de Atkinson, 2025); d, vista lateral de C. hispidus (tomada de Atkinson, 2025); e, vista frontal de C. squamifer (tomada de Atkinson, 2025); f, vista frontal de C. hispidus (tomada de Atkinson, 2025).
Subtribu Corthylina LeConte, 1876
Funículo antenal con 1-5 segmentos, maza grande y generalmente asimétrica; pieza intercoxal prosternal ausente, pubescencia comúnmente muy reducida, es insignificante muy confusa (no ordenada en filas); el declive elitral es convexo o truncado a profundamente excavado, generalmente con procesos similares a espinas. La mayoría de las especies perforan los tejidos del xilema del huésped e inoculan esporas de sus hongos simbióticos en las paredes del túnel, luego se alimentan principalmente de las esporas del hongo y el micelio (Wood, 2007).
Clave de las especies de la subtribu Corthylina
1. Funículo antenal de 2 segmentos; masa antenal truncada apicalmente, ovalada o redonda 2
1’. Funículo antenal de 1 segmento; masa antenal ovalada o redonda 3
2. Cuerpo castaño con 3 manchas más claras y rojizas en base del pronoto y 2 manchas alargadas longitudinales amarillentas en élitros, van desde la base hasta antes del declive elitral (fig. 5a, c); pronoto 1.5 veces más largo que ancho, borde anterior del pronoto armado con dientes; declive elitral amplio, moderadamente cóncavo, espina 1 en interestría 1 diminuta y puntiaguda, 2 en interestría 2, ligeramente más grandes que 1, espina 3 diminuta, espina 4 ligeramente más grandes que 2, puntiagudas (fig. 5e Monarthrum chapuisi Kirsch
2’. Cuerpo castaño rojizo muy oscuro y homogéneo (fig. 5b, d); pronoto 1.3 veces más largo que ancho, borde anterior del pronoto liso (sin dientes); declive elitral empinado, reticulado, estrecho y suavemente impreso, con espinas 2 y 3 diminutas (fig. 5f Monarthrum subimpressum Wood
3. Maza antenal 1.5 veces más larga que ancha, con sutura visible y cirro delgado (fig. 6e); frente fuertemente cóncava, mitad inferior amarillenta y esponjosa, sin setas (fig. 6g); margen anterior del pronoto débilmente aserrado; declive elitral muy empinado, redondeado y convexo, con setas erectas largas difusas (fig. 6c) Corthylus alineus Schedl
3’. Maza antenal 1.3 veces más larga que ancha, con 3 suturas visibles y cirro muy largo (fig. 6f); frente moderadamente cóncava, con áreas laterales a ambos lados de la frente, amarillas y esponjosas, con setas largas y erectas a ambos lados del área esponjosa (fig. 6h); borde apical de la frente con fila de setas plumosas orientadas hacia el centro de la frente; margen anterior del pronoto con espinas; declive elitral muy empinado (casi truncado) débilmente convexo, con cresta costal elevada no cerrada, con escasas setas largas y erectas (fig. 6d) Corthylus serrulatus Eggers
Monarthrum chapuisi Kirsch, 1866 (fig. 3a, c, e)
Distribución. Argentina: Tucumán; América del Sur: Bolivia, Colombia, Perú y Venezuela (Atkinson, 2025; Córdoba et al., 2023, 2025; Smith et al., 2017; Wood, 2007).
Monarthrum subimpressum Wood, 2007 (fig. 5b, d, f)
Distribución. Argentina: Salta, Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Corthylus alineus Schedl, 1966 (fig. 6a, c, e, g)
Distribución. Argentina: Salta y Tucumán; América del Sur: Colombia, Ecuador y Perú (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2023, 2025; Smith et al., 2017; Wood, 2007).
Corthylus serrulatus Eggers, 1934 (fig. 6b, d, f, h)
Distribución. Argentina: Jujuy, Salta y Tucumán; América del Sur: Bolivia, Brasil y Perú (Atkinson, 2025; Córdoba et al., 2018, 2021, 2023, 2025; Smith et al., 2017; Wood, 2007).
Subtribu Pityophthorina Eichhoff, 1878
Se caracteriza por presentar el funículo antenal mayormente de 5 segmentos, maza antenal más pequeña y simétrica; pieza intercoxal prosternal agudamente puntiaguda; pubescencia abundante, generalmente en filas en los élitros; declive elitral comúnmente convexo a bisulcado, armadura ausente. Se alimentan directamente del tejido de la planta huésped y son floeófagas, mielófagas o espermófagas. No hay especies verdaderamente xilófagas (Wood, 2007).
Clave de las especies de la subtribu Pityophthorina
1. Revestimiento del cuerpo formado por escamas (fig. 7a, b); flagelo antenal con 2 segmentos; maza antenal alargada y subrectangular (fig. 7c); pronoto con 8 filas transversales de espinas romas; ramillete de setas largas en cara interna de protibias, que van desde la base hasta el ápice de las mismas (fig. 7e); menos de 2 mm de longitud del cuerpo Acorthylus bosqui (Schedl)
1’. Revestimiento del cuerpo formado por setas; flagelo antenal con 4 segmentos; maza antenal redondeada y aplanada (fig. 7d); pronoto con espinas romas o sin espinas; tibias anteriores con pubescencia abundante (fig. 7f); más de 2 mm de longitud del cuerpo 2
2. Pronoto con rugosidades pronunciadas en la parte anterior; pronoto en vista lateral, convexo hasta el disco, luego con una elevación pronunciada en medio, más bajo y recto hasta la base; maza antenal con 2 suturas parcial o totalmente septadas 3
2’. Pronoto sin rugosidades pronunciadas en la parte anterior; pronoto en vista lateral con curvatura uniforme desde el ápice hasta la base, sin elevación pronunciada; maza antenal con 1 sutura total o parcialmente septada 5
3. Declive elitral con sulcos, interestría 1 moderadamente elevada, armada por 10 o más tubérculos; pubescencia de los élitros confinada al declive (fig. 9e) Pityophthorus tucumanensis Wood
3’. Declive elitral completamente convexo, sin sulcos, interestría 1 sin tubérculos 4
4. Longitud del cuerpo 1.4 a 1.8 mm; frente con fuerte impresión transversal y con pequeño tubérculo medio a nivel de los ojos (fig. 8e); color del cuerpo castaño oscuro (fig. 8a, c) Cryptocarenus heveae (Hagedorn)
4’. Longitud del cuerpo 1.4 a 1.5 mm; frente cóncava (fig. 8f); coloración del cuerpo castaño rojizo o claro (fig. 8b, d) Cryptocarenus seriatus Eggers
5. Pronoto de lados paralelos; élitros 2 veces más largos que el pronoto (fig. 9b, d) Araptus araujiae (Brèthes)
5’. Pronoto de lados convexos; élitros entre 1.6 y 1.7 más largos que el pronoto 6
6. Frente con quilla media aguda, más fuertemente elevada en mitad media superior (fig. 10g); maza antenal redondeada; pubescencia del cuerpo abundante (fig. 10a,d) Araptus pubescens (Schedl)
6’. Frente sin quilla; maza antenal ovalada; pubescencia no muy abundante 7
7. Frente muy débilmente convexa, con pubescencia corta y escasa (fig. 10h); declive elitral convexo y bisulcado (fig. 10j) Araptus frenatus (Schedl)
7’. Frente claramente convexa, con pubescencia larga abundante y erecta; con tubérculo en el ápice de la frente, poco marcado que se afina hasta la base (fig. 10i); declive elitral convexo con surco poco profundo en las interestrías 1 y 2 (fig. 10k) Araptus volastos (Schedl)
Figura 5. a, Vista dorsal de Monarthrum chapuisi (tomada de Atkinson, 2025); b, vista dorsal de M. subimpresum (tomada de Atkinson, 2025); c, vista lateral de M. chapuisi (tomada de Atkinson, 2025); d, vista lateral de M. subimpresum (tomada de Atkinson, 2025); e, vista posterior de M. chapuisi; f, vista posterior de M. subimpresum (tomada de Atkinson, 2025).
Figura 6. a, Vista dorsal de Corthylus alineus;b, vista dorsal de C. serrulatus (tomada de Atkinson, 2025); c, vista lateral de C. alineus; d, vista lateral de C. serrulatus (tomada de Atkinson, 2025); e, antena de C. alineus; f, antena de C. serrulatus; g, vista frontal de C. alineus; h, vista frontal de C. serrulatus.
Acorthylus bosqui (Schedl, 1938) (fig. 7a, b, c, e)
Distribución. Argentina: Jujuy y Tucumán; América del Sur: Bolivia (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Cryptocarenus heveae (Hagedorn, 1912) (fig. 7d, f, 8a, c, e)
Distribución. Argentina: Buenos Aires, Misiones, Santiago del Estero y Tucumán; América del Sur: Brasil, Colombia Perú, Trinidad y Tobago y Venezuela (Atkinson, 2025; Córdoba et al., 2021, 2023, 2025; Iturre y Darchuck, 1996; Smith et al., 2017; Wood, 2007).
Cryptocarenus seriatus Eggers, 1933 (fig. 8b, d, f)
Distribución. Argentina: Tucumán; América del Sur: Bolivia, Brasil, Colombia, Guyana Francesa, Paraguay, Perú y Venezuela (Atkinson 2025; Córdoba et al., 2023, 2025; Smith et al., 2017; Wood, 2007).
Pityophthorus tucumanensis Wood, 2007 (fig. 9a, c, e)
Distribución. Argentina: Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Araptus araujiae (Brèthes, 1921) (fig. 9b, d)
Distribución. Argentina: Buenos Aires y Tucumán (Atkinson, 2025; Bachmann y Lanteri, 2013; Córdoba et al., 2023, 2025; Wood y Bright, 1992).
Araptus pubescens (Schedl, 1950) (fig. 10a, d, g)
Distribución. Argentina: Córdoba y Tucumán (Atkinson, 2025; Córdoba et al., 2025; Wood, 2007).
Araptus frenatus (Schedl, 1939) (fig. 10b, e, h, j)
Distribución. Argentina: Córdoba y Tucumán (Atkinson, 2025; Córdoba et al., 2025; Wood, 2007).
Araptus volastos (Schedl, 1938) (fig. 10c, f, i, k)
Distribución. Argentina: Jujuy, Salta y Tucumán; América del Sur: Bolivia (Atkinson, 2025; Córdoba et al., 2025; Wood, 2007).
Tribu Dryocoetini Lindemann, 1877
Figura 7. a, Vista dorsal de Acorthylus bosqui (tomada de Atkinson, 2025); b, vista lateral de A. bosqui (tomada de Atkinson, 2025); c, antena de A. bosqui; d, antena de Cryptocarenus heveae; e, tibia de A. bosqui; f, tibia de C. heveae (tomada de Atkinson, 2025).
Figura 8. a, Vista dorsal de Cryptocarenus heveae (tomada de Atkinson, 2025); b, vista dorsal de C. seriatus (tomada de Atkinson, 2025); c, vista lateral de C. heveae (tomada de Atkinson, 2025); d, vista lateral de C. seriatus (tomada de Atkinson, 2025); e, vista frontal de C. heveae (tomada de Atkinson, 2025); f, vista frontal de C. seriatus
Presentan dimorfismo sexual en la frente. En los machos es convexa a variadamente impresa y en las hembras es convexa a aplanada o con elevaciones; el ojo está emarginado o dividido; el escapo antenal es alargado, el funículo está compuesto de 4 a 6 segmentos, la maza antenal puede estar oblicuamente truncada o conspicuamente aplanada, en este último caso las suturas son procurvadas u obsoletas; el pronoto puede estar o no armado con asperezas; las procoxas pueden ser contiguas o estar muy poco separadas; margen lateral de las protibias con 3 o más dentículos alveolares; el declive elitral puede ser convexo, sulcado, aplanado y, a veces, con pequeños gránulos; la pubescencia está compuesta por setas; la mayoría de las especies son polígamas y algunas son haploides; hay especies espermatófagas, mielófagas y floeófagas. De los 17 géneros que se conocen en el mundo, solo 4 se encuentran en América del Sur, pero 2 de ellos son introducidos (Wood, 2007). Las especies del género Coccotrypes presentes en la región son barrenadores de semillas de palmeras.
Clave de las especies de la tribu Dryocoetini
1. Tamaño del cuerpo entre 1.8 a 2.3 mm; pubescencia moderadamente abundante (fig. 11a, c); pronoto de lados marcadamente convexos que se angosta de manera abrupta hacia el ápice, con asperezas pequeñas y abundantes (fig. 11a) Coccotrypes dactyliperda (Fabricius)
1’. Tamaño del cuerpo entre 1.5 y 1.9 mm; pubescencia muy abundante (fig. 11b, d); pronoto redondeado, angostado en el ápice, con asperezas gruesas y marcadas (fig. 11b) Coccotrypes carpophagus (Hornung)
Coccotrypes dactyliperda (Fabricius, 1801)* (fig. 11a, c)
Distribución. S/D, Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood y Bright, 1992).
Coccotrypes carpophagus (Hornung, 1842)* (fig. 11b, d)
Distribución. Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025).
Tribu Hexacolini Eichhoff, 1878
Sus miembros presentan dimorfismo sexual: los machos pueden presentar una impresión en la frente, a veces oscura; los machos son convexos y las hembras son esculpidas y ornamentadas de forma variada; los ojos son en general alargados, con su margen anterior entero o sinuoso; el escapo antenal es alargado; el funículo lleva 5 o 6 segmentos (con 7 segmentos en Gymnochilus); la maza antenal puede o no llevar suturas; la zona apical del pronoto puede estar ornamentada o no; las procoxas están ampliamente separadas; las protibias llevan, en el margen lateral, 1 o más dentículos alveolares generalmente incrustados en la cutícula (Wood, 2007).
Clave de las especies de la tribu Hexacolini
1. Color castaño amarillento, con franja más oscura en mitad apical del pronoto (fig. 12a, c); frente con callo transversal por debajo del nivel superior de ojos (fig. 12e); pronoto con superficie reticulada; élitros de lados convexos (fig. 12a); declive elitral ampliamente convexo (fig.12c) Scolytodes sparsepilosus Wood
1’. Color castaño amarillento, con mitad apical más oscura (fig. 12b, d); frente con callo medio estrecho que se extiende dorsalmente desde elevación epistomal hasta ligeramente por encima del nivel de inserción de antena (fig. 12f); pronoto liso y brillante; élitros de lados paralelos (fig. 12b); declive estrechamente convexo (fig. 12d) Scolytodes tucumani Wood
Scolytodes sparsepilosus Wood, 2007 (fig. 12a, c, e)
Distribución. Argentina: Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Scolytodes tucumani Wood, 2007 (fig. 12b, d, f)
Distribución. Argentina: Tucumán; América del Sur: Ecuador (Atkinson, 2025; Córdoba et al., 2023, 2025; Jordal y Smith, 2020; Wood, 2007).
Tribu Hylurgini Gistel, 1848
Presentan dimorfismo sexual en la frente, siendo convexa en las hembras e impresa en los machos; los ojos son enteros y ovalados; las antenas presentan un escapo alargado, el funículo con 5 a 7 segmentos y la maza antenal es simétrica, aplanada y generalmente con 3 o 4 suturas; el pronoto es generalmente liso; las procoxas son contiguas o están muy poco separadas; las tibias están armadas por dientes en los márgenes laterales y apicales; todas las especies de esta tribu son monógamas y fleófagas (Wood, 2007).
Figura 9. a, Vista dorsal de Pityophthorus tucumanensis (tomada de Atkinson, 2025); b, vista dorsal de Araptus araujiae; c, vista lateral de P. tucumanensis (tomada de Atkinson, 2025); d, vista lateral de A. araujiae; e, vista posterior de P. tucumanensis (tomada de Atkinson, 2025).
Clave de las especies de la tribu Hylurgini
Longitud 2.1-2.6 mm; color castaño oscuro, con pubescencia formada por setas y escamas con patrón de coloración mezclada (fig. 13a, b); frente convexa, reticulada y rugosa por encima de los ojos y lisa y brillante por debajo de ellos (fig. 13c); pronoto 0.85 veces más largo que ancho, base más ancha que el ápice, lados convexos, con superficie reticulada (fig. 13a); interestrías de élitros armadas por tubérculos; en macho las interestrías 3 y base de la 1 elevadas y armadas por dentículos en declive elitral Xylechinus imperialis (Schedl)
Xylechinus imperialis (Schedl, 1958) (fig. 13a-c)
Distribución. Argentina: Buenos Aires, Jujuy, Salta, Santa Fe y Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Tribu Ipini Bedel, 1888
Se distinguen por presentar dimorfismo sexual en la frente, siendo convexa en los machos y de forma variada en las hembras; la antena presenta un escapo afinado y alargado, el funículo antenal está formado por 5 segmentos y la maza puede ser oblicuamente truncada o aplanada, con suturas desplazadas hacia el ápice de la cara posterior; mitad anterior del ápice del pronoto es declivado y áspero; las procoxas son contiguas; protibias presentan 3 o 4 dientes; el declive elitral es sulcado o profundamente excavado, puede estar armado con espinas o tubérculos; el cuerpo presenta una pubescencia similar a setas. Todas las especies de esta tribu son floeófagas y monógamas o polígamas (Wood, 2007).
Figura 10. a, Vista dorsal de Araptus pubescens (tomada de Atkinson, 2025); b, vista dorsal de A. frenatus (tomada de Atkinson, 2025); c, vista dorsal de A. volastos (tomada de Atkinson, 2025); d, vista lateral de A. pubescens (tomada de Atkinson, 2025); e, vista lateral de A. frenatus (tomada de Atkinson, 2025); f: vista lateral de A. volastos (tomada de Atkinson, 2025); g: vista frontal de A.pubescens; h, vista frontal de A. frenatus (tomada de Atkinson, 2025); i, vista frontal de A. volastos; j, vista posterior de A. frenatus (tomada de Atkinson, 2025); k, vista posterior de A. volastos (tomada de Atkinson, 2025).
Figura 11. a, Vista dorsal de Coccotrypes dactyliperda (tomada de Atkinson, 2025); b, vista dorsal de C. carpophagus (tomada de Atkinson, 2025); c, vista lateral de C. dactyliperda (tomada de Atkinson, 2025); d, vista lateral de C. carpophagus (tomada de Atkinson, 2025).
Figura 12. a, Vista dorsal de Scolytodes sparsepilosus (tomada de Atkinson, 2025); b, vista dorsal de S. tucumani (tomada de Atkinson, 2025); c, vista lateral de S. sparsepilosus (tomada de Atkinson, 2025); d, vista lateral de S. tucumani (tomada de Atkinson, 2025); e, vista frontal de S. sparsepilosus (tomada de Atkinson, 2025); f, vista frontal de S. tucumani (tomada de Atkinson, 2025).
Clave de las especies de la tribu Ipini
(Macho) Longitud: 3.0-3.7 mm; 2.7 veces más largo que ancho (fig. 14a, b); frente levemente convexa con abundante pubescencia larga, fina y erecta; pronoto 1,07 veces más largo que ancho, mitad anterior moderadamente declinada y mitad basal levemente cóncava, con pubescencia escasa en el centro y con setas largas y erectas en los bordes y ápice (fig. 14b); declive elitral cóncavo, con 3 espinas en bordes externos, 1 espina en interestría 1 y otra en la interestría 2 (fig. 14c) Orthotomicus laricis (Fabricius)
Orthotomicus laricis (Fabricius, 1792)* (fig. 14a, b, c)
Distribución. Argentina: Neuquén y Tucumán; América del Sur: Chile (Atkinson, 2025; Córdoba et al., 2021, 2023, 2025; Kirkendall, 2018; Wood, 2007).
Tribu Micracidini LeConte, 1876
Se distinguen por presentar dimorfismo sexual en la frente, siendo en la hembra generalmente cóncava y en el macho, raramente cóncava; los ojos pueden ser ovalados, alargados, enteros o sinuados; antenas con el escapo corto o alargado, aplanado o expandido, con setas, el funículo presenta 6 segmentos, la maza antenal puede o no tener suturas; el pronoto presenta asperezas en la zona anterior, los lados son redondeados; las procoxas están ligeramente separadas; generalmente presentan setas subplumosas en alguna parte del cuerpo; pueden tener hábitos xilófagos, floeófagos o mielófagos; todos son bígamos excepto el género Micraciella que es monógamo (Wood, 2007).
Clave de las especies de la tribu Micracidini
(Macho) Longitud 3.2 a 3.5 mm; 2.8 veces más largo que ancho; frente convexa con carena transversal (fig. 15c); antena con escapo ligeramente aplanado con mechón de pelos (fig. 15c), sutura procurvada; mitad apical del pronoto con tubérculos que van disminuyendo hacia la mitad, margen apical con 6 dientes; declive elitral convexo con nódulos en las interestrías de la base, interestría 9 con quilla, pubescencia escasa, aumenta en el declive elitral, élitros terminados en punta aguda (fig. 15d) Hylocurus giganteus (Schedl)
Hylocurus giganteus (Schedl, 1950) (fig. 15a-d)
Distribución. Argentina: Salta y Tucumán; América del Sur: Brasil (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Tribu Phloeosinini Nüsslin, 1912
Presentan dimorfismo sexual en la frente, machos con diferentes impresiones y hembras con frente aplanada o convexa; ojos pueden estar emarginados; antenas llevan un funículo con 5-7 segmentos, maza antenal aplanada, pueden ser asimétricas en diferentes medidas, pueden llevar suturas o no; pronoto no tiene asperezas, excepto en Dendrosinus, que presenta asperezas débiles; tercer segmento tarsal comprimido o bilobulado; escutelo puede o no ser visible; pueden ser polígamos o bígamos; pueden ser floeófagos o xilófagos (Wood, 2007).
Clave para las especies de la tribu Phloeosinini
1. Maza antenal fuertememente aplanada, en forma de riñón, la unión con el funículo es en su parte lateral, sin suturas ni fila de setas; escutelo presente pero diminuto; pronoto liso o reticulado 2
1’. Maza antenal con suturas fuertemente procurvadas y claramente marcadas por fila de setas, unión con el funículo excéntrica; escutelo visible; pronoto liso o armado con pequeños gránulos 4
2. Frente del macho con márgenes laterales fuertemente elevados y con cuentas 3
2’. Frente del macho con márgenes laterales agudamente elevados con cresta irregular pero sin cuentas (fig. 16g) Chramesus phloeotriboides Schedl
3. Margen lateral de la frente del macho fuertemente elevado y con cuentas solo en el tercio inferior (fig. 16h); maza antenal 2.6 veces más larga que ancha; pronoto con superficie lisa y brillante, zona basal con algunos tubérculos pequeños y aislados Chramesus argentinae Wood
3’. Margen lateral de la frente del macho agudamente elevado y marcado con 9 cuentas (fig. 16i); maza antenal 2.3 veces más larga que ancha; pronoto con zona basal profundamente punteada Chramesus globosus Hagedorn
4. Frente ampliamente convexa; estrías de los élitros claramente impresas o no impresas con punciones regulares 5
4’. Frente ancha y ligeramente cóncava (fig. 17g); estrías de los élitros impresas con punciones bastantes grandes; interestría 7 elevada en los machos Pseudochramesus costulatus Blackman
5. Interestrías con escamas de fondo castañas oscuras y pálidas, formando un patrón, setas pálidas uniseriadas en interestrías (fig. 17b, e); frente con elevación a nivel de inserción de antenas (fig. 17h) Pseudochramesus acuteclavatus (Hagedorn)
5’. Interestrías con escamas de fondo castañas, a lo largo de la sutura, pálidas (fig. 17c, f); frente sin elevación a nivel de la inserción de las antenas (fig. 17i) Pseudochramesus harringtoni Blackmann
Chramesus phloeotriboides Schedl, 1958 (fig. 16a, d, g)
Distribución. Argentina: Córdoba y Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Chramesus argentinae Wood, 2007 (fig. 16b, e, h)
Distribución. Argentina: Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Chramesus globosus Hagedorn, 1909 (fig. 16c, f, i)
Distribución. Argentina: Buenos Aires, La Rioja, Misiones, Santa Fe y Tucumán; América del Sur: Brasil y Uruguay (Atkinson, 2025; Bachmann y Lanteri, 2013; Córdoba et al., 2023, 2025; Wood, 2007).
Pseudochramesus costulatus Blackman, 1939 (fig. 17a, d, g)
Distribución. Argentina: Jujuy y Tucumán; América del Sur: Bolivia (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Pseudochramesus acuteclavatus (Hagedorn, 1909) (fig. 17b, e, h)
Distribución. Argentina: Buenos Aires, Salta y Tucumán; América del Sur: Bolivia, Brasil y Paraguay (Atkinson, 2025; Bachmann y Lanteri, 2013; Córdoba et al., 2023, 2025; Wood, 2007).
Pseudochramesus harringtoni Blackmann, 1939 (fig. 17c, f, i)
Distribución. Argentina: Salta y Tucumán; América del Sur: Bolivia y Brasil (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Tribu Phloeotribini Chapuis, 1869
Presentan dimorfismo sexual en la frente, siendo con impresiones variadas en los machos y planas o convexas en las hembras; los ojos son enteros; antenas con funículo de 5 segmentos, la maza antenal puede ser muy delgada o fuertemente asimétrica, está dividida en 3 segmentos móviles, usualmente sublamelados; las procoxas son contiguas; el pronoto puede llevar o no asperezas, los márgenes laterales son redondeados; son especies monógamas y floeófagas (Wood, 2007).
Figura 13. a, Vista dorsal de Xylechinus imperialis (tomada de Atkinson, 2025); b, vista lateral de X. imperialis (tomada de Atkinson, 2025); c, vista frontal de X. imperialis (tomada de Atkinson, 2025).
Figura 14. a, Vista dorsal de Orthotomicus laricis (tomada de Atkinson, 2025); b, vista lateral de O. laricis (tomada de Atkinson, 2025); c, vista frontal de O. laricis (tomada de Atkinson, 2025).
Figura 15. a, Vista dorsal de Hylocurus giganteus (tomada de Atkinson, 2025); b, vista lateral de H. giganteus (tomada de Atkinson, 2025); c, vista frontal de H. giganteus; d, vista posterior de H. giganteus (tomada de Atkinson, 2025).
Figura 16. a, Vista dorsal de Chramesus phloeotriboides (tomada de Atkinson, 2025); b, vista dorsal de C. argentinae (tomada de Atkinson, 2025); c, vista dorsal de C. globosus (tomada de Atkinson, 2025); d, vista lateral de C. phloeotriboides (tomada de Atkinson, 2025); e, vista lateral de C. argentinae (tomada de Atkinson, 2025); f, vista lateral de C. globosus (tomada de Atkinson, 2025); g, vista frontal de C. phloeotriboides (tomada de Atkinson, 2025); h, vista frontal de C. argentinae (tomada de Atkinson, 2025); i, vista frontal de C. globosus (tomada de Atkinson, 2025).
Figura 17. a, Vista dorsal de Pseudochramesus costulatus (tomada de Atkinson, 2025); b, vista dorsal de P. acuteclavatus (tomada de Atkinson, 2025); c, vista dorsal de P. harringtoni (tomada de Atkinson, 2025); d, vista lateral de P. costulatus (tomada de Atkinson, 2025); e, vista lateral de P. acuteclavatus (tomada de Atkinson, 2025); f, vista lateral de P. harringtoni (tomada de Atkinson, 2025); g, vista frontal de P. costulatus (tomada de Atkinson, 2025); h, vista frontal de P. acuteclavatus (tomada de Atkinson, 2025); i, vista frontal de P. harringtoni (tomada de Atkinson, 2025).
Clave para las especies de la tribu Phloeotribini
1. Frente con área cóncava; pronoto con crenulaciones en mitad apical y puntos grandes; élitros con interestrías iguales o más angostas que estrías; declive elitral con espinas 2
1’. Frente con área cóncava y carena epitosmal muy elevada en macho (fig. 18g); pronoto con asperezas en mitad apical y puntos medianos; élitros con interestrías 3 veces más anchas que estrías; declive elitral con tubérculos pequeños (fig. 18j) Phloeotribus subovatus Blandford
2. Escapo antenal con mechón grande de setas largas (fig. 18h); interestrías de igual ancho que estrías; declive elitral con tubérculos espiniformes en interestrías 2-9 (fig. 18k) Phloeotribus asperulus Eggers
2’. Escapo antenal con escasas setas (fig. 18i); interestrías más angostas que las estrías; declive elitral con espinas en las interestrías 1 a la 9 (fig. 18l) Phloeotribus harringtoni Blackman
Phloeotribus subovatus Blandford, 1897 (fig. 18a, d, g, j)
Distribución. Argentina: Jujuy, Salta y Tucumán; América del Sur: Perú y Venezuela (Atkinson, 2025; Córdoba et al., 2025; Wood, 2007).
Phloeotribus asperulus Eggers, 1943 (fig. 18b, e, h, k)
Distribución. Argentina: Tucumán; América del Sur:
Bolivia y Brasil (Atkinson, 2025; Córdoba et al., 2025; Wood y Bright, 1992).
Phloeotribus harringtoni Blackman, 1943 (fig. 18c, f, i, l)
Distribución. Argentina: Salta y Tucumán (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Tribu Scolytini Latreille, 1804
Figura 18. a, Vista dorsal de Phloeotribus subovatus (tomada de Atkinson, 2025); b, vista dorsal de P. asperulus (tomada de Atkinson, 2025); c, vista dorsal de P. harringtoni (tomada de Atkinson, 2025); d, vista lateral de P. subovatus (tomada de Atkinson, 2025); e, vista lateral de P. asperulus (tomada de Atkinson, 2025); f, vista lateral de P. harringtoni (tomada de Atkinson, 2025); g, vista frontal de P. subovatus; h, vista frontal de P. asperulus (tomada de Atkinson, 2025); i, vista frontal de P. harringtoni (tomada de Atkinson, 2025); j, vista posterior de P. subovatus (tomada de Atkinson, 2025); k, vista posterior de P. asperulus (tomada de Atkinson, 2025); l, vista posterior de P. harringtoni (tomada de Atkinson, 2025).
Protibias y, usualmente, metatibias desarmadas de espinas, solo llevan un único proceso similar a una espina curvado en el ángulo apical lateral; los márgenes laterales del pronoto son subagudos y elevados; antena con funículo de 7 segmentos, maza antenal puede presentar suturas fuertemente procurvadas, parciales u obsoletas; frente en los machos puede llevar impresiones y la de las hembras es convexa; los ojos son ovalados y enteros (Wood, 2007).
Clave para las especies de la tribu Scolytini
1. Macho. Frente ampliamente convexa, con setas finas moderadamente abundantes (fig. 19g); pronoto 1 vez más largo que ancho; estrías e interestrías confusas; abdomen gradualmente ascendente hasta los élitros (fig. 19d) Scolytus rugulosus (Müller)
1’. Macho. Frente oculta por cepillo de setas largas (fig. 19h, i); pronoto tan largo como ancho; abdomen abruptamente flexionado hacia arriba desde el margen posterior del esternito visible 2 (fig. 19e, f) 2
2. Pronoto con puntos pequeños en el disco que aumentan de tamaño hacia las zonas laterales; setas interestriales robustas y 10 veces más largas que anchas Scolytopsis toba Wichmann
2’. Pronoto con puntos medianos en el disco y con zona anterior y laterales reticulada; setas interestriales 4 veces más largas que anchas Scolytopsis punticollis Blandford
Scolytus rugulosus (Müller, 1818)* (fig. 19a, d, g)
Distribución. Buenos Aires, Catamarca, La Rioja, Mendoza y Misiones; América del Sur: Brasil, Chile, Perú y Uruguay (Atkinson, 2025; Smith et al., 2017; Córdoba y Atkinson, 2018; Córdoba et al., 2023, 2025; Wood, 2007).
Comentario. Esta especie es de origen euroasiático, pero se distribuye actualmente en todas las regiones templadas del mundo.
Scolytopsis toba Wichmann, 1914 (fig. 19b, e, h)
Distribución. Argentina: Misiones y Tucumán; América del Sur: Brasil y Paraguay (Atkinson, 2025; Córdoba et al., 2025; Petrov, 2017; Wood, 2007).
Scolytopsis punticollis Blandford, 1896 (fig. 19c, f, i)
Distribución. Argentina: Misiones y Tucumán; América del Sur: Brasil (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Tribu Trypophloeini Nüsslin, 1911
Figura 19. a, Vista dorsal de Scolytus rugulosus (tomada de Atkinson, 2025); b, vista dorsal de Scolytopsis toba (tomada de Atkinson, 2025); c, vista dorsal de S. punticollis (tomada de Atkinson, 2025); d, vista lateral de S. rugulosus (tomada de Atkinson, 2025); e, vista lateral de S. toba (tomada de Atkinson, 2025); f, vista lateral de S. punticollis (tomada de Atkinson, 2025); g, vista frontal de S. rugulosus (tomada de Atkinson, 2025); h, vista frontal de S. toba (tomada de Atkinson, 2025); i, vista frontal de S. punticollis (tomada de Atkinson, 2025).
Se diferencian por la presencia del tercer tarso de forma cilíndrica; ojos son emarginados; funículo antenal con 3-5 segmentos y maza puede presentar suturas y un único septo parcial; pubescencia del hipomeron está compuesta de setas simples raramente mezclado con setas bifurcadas; setas de interestrías similares a escamas; macho similar a la hembra, de tamaño más pequeño y no volador en Hypothenemus, generalmente hay dimorfismo sexual en la frente (Johnson et al., 2020).
Figura 20. a, Vista dorsal de Hypothenemus meridensis (tomada de Atkinson, 2025); b, vista dorsal de H. eruditus (tomada de Atkinson, 2025); c, vista lateral de H. meridensis (tomada de Atkinson, 2025); d, vista lateral de H. eruditus (tomada de Atkinson, 2025); e, vista dorsal del pronoto de H. meridensis; f, vista dorsal del pronoto de H. eruditus.
Clave para las especies de la tribu Trypophloeini
1. Margen anterior del pronoto con 2 dientes (fig. 20e); pendiente anterior del pronoto con dientes gruesos elevados Hypothenemus meridensis,Wood
1’. Margen anterior del pronoto con 6 dientes (fig. 20f); pendiente anterior del pronoto con dientes medianos 2
2. Tamaño pequeño, 1.0- 1.3 mm; frente con superficie rugosa y reticulada; élitros con interestrías 2 veces más anchas que estrías Hypothenemus eruditus Westwood
2’. Tamaño del cuerpo de 1.4- 1.6 mm; frente rugosa y reticulada, con surco o tubérculo; élitros con estrías tan anchas como interestrías 3
3. Frente con tubérculo pequeño ubicado encima del nivel superior de ojos y surco poco profundo que se extiende desde el tubérculo hasta el epistoma (fig. 21e); declive elitral fuertemente convexo (fig. 21c) Hypothenemus crudiae (Panzer)
3’. Frente con surco poco profundo que llega hasta la tercera parte de distancia del margen del epistoma (fig. 21f); declive elitral convexo (fig. 21d) Hypothenemus seriatus (Eichhoff)
Hypothenemus meridensis Wood, 2007 (fig. 20a, c, e)
Distribución. Argentina: Tucumán; América del Sur: Brasil y Venezuela (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Hypothenemus eruditus Westwood, 1836 (fig. 20b, d, f)
Distribución. Argentina: Buenos Aires, Corrientes, Misiones, Santiago del Estero y Tucumán; América del Sur: Brasil, Colombia, Ecuador, Guyana, Trinidad y Tobago y Venezuela (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Smith et al., 2017; Wood, 2007).
Hypothenemus crudiae (Panzer, 1791)* (fig. 21a, c, e)
Distribución. Argentina: Buenos Aires y Tucumán; América del Sur: Bolivia, Brasil, Colombia, Ecuador, Guyana, Paraguay, Surinam, Trinidad y Tobago y Venezuela (Atkinson, 2025; Atkinson y Flechtmann, 2021; Córdoba et al., 2023, 2025; Wood, 2007).
Comentarios. Esta especie se distribuye desde Estados Unidos hasta Argentina, también posiblemente introducida en Asia y África.
Hypothenemus seriatus (Eichhoff, 1872) (fig. 21b, d, f)
Distribución. Argentina: Tucumán; América del Sur: Bolivia, Brasil, Colombia, Ecuador, Paraguay, Perú y Venezuela (Atkinson, 2025; Córdoba et al., 2021, 2023, 2025; Smith et al., 2017; Wood, 2007).
Tribu Xyleborini LeConte, 1876
Figura 21. a, Vista dorsal de Hypothenemus crudiae (tomada de Atkinson, 2025); b, vista dorsal de H. seriatus (tomada de Atkinson, 2025); c, vista lateral de H. crudiae (tomada de Atkinson, 2025); d, vista lateral de H. seriatus (tomada de Atkinson, 2025); e, vista frontal de H. crudiae (tomada de Atkinson, 2025); f, vista frontal de H. seriatus.
Se caracterizan por presentar dimorfismo sexual marcado, macho generalmente más pequeño que hembra y no volador, ojo reducido en tamaño y haploide, la hembra normal y diploide; frente convexa, sin ornamentaciones; ojos son emarginados o divididos; frecuentemente escapo antenal alargado, funículo lleva 5 segmentos y maza antenal truncada de manera oblicua; parte anterior del pronoto empinada y armada con asperezas; procoxas varían de contiguas a ampliamente separadas; protibias expandidas, arqueadas y armadas con dentículos; escutelo puede ser grande y plano o modificado o ausente; en cuanto a su biología, la xilomicetofagia y la poligamia endogámica son universales en esta tribu (Wood, 2007).
Clave para las especies de la tribu Xyleborini
1. Declive de élitros abruptamente truncados con quilla circundeclivital, cara del declive ligeramente convexa con 3 estrías marcadas (fig. 22a, c, e); élitros tan largos como pronoto; procoxas contiguas (fig. 22g) Amasa parviseta Knížek et Smith
1’. Declive elitral aplanado o convexo, nunca truncado; élitros más largos que pronoto; procoxas contiguas o separadas 2
2. Escutelo en forma de espina y poco visible, dejando base de élitros con sutura hueca y cubierta de setas 3
2’. Escutelo en forma triangular y visible, concavidad sutural completamente llena hasta uperficie anterior de élitros, emarginación sin setas 5
3. Declive elitral convexo, empinado y débilmente impreso (fig. 22d, f); margen anterior del pronoto armado por dientes pequeños; longitud del cuerpo 1.9 a 2.4 mm Xyleborinus saxesenii (Ratzeburg)
3’. Declive elitral empinado e impreso; margen anterior del pronoto débilmente aserrado; longitud del cuerpo 2.0 a 2.6 mm 4
4. Frente reticulada, toscamente punteada (fig. 23e); longitud del cuerpo 2.0 a 2.2 mm y 2.9 más largo que ancho; declive elitral empinado e impreso hasta la sutura 3, con interestrías 1 y 2 impresas sin tubérculos, 3 débilmente elevada con 4 espinas (3 de mayor tamaño), intercaladas con otras más pequeñas, interestría 4 con tubérculos puntiagudos pequeños (fig. 23g) Xyleborinus linearicollis (Schedl)
4’. Frente reticulada con punciones gruesas y profundas (fig. 23f); longitud del cuerpo 2.3 a 2.6 mm y 3.2 veces más largo que ancho; declive elitral corto, muy empinado, fuertemente impreso desde la sutura hasta la interestría 3, interestría 1 con un tubérculo en la base, interestría 3 moderadamente elevada con 3 espinas grandes igualmente espaciadas (fig. 23h) Xyleborinus sentosus (Eichhoff)
5. Procoxas moderada o ampliamente separadas; pieza intercoxal continua, longitudinalmente sin emarginación 6
5’. Procoxas contiguas; pieza intercoxal longitudinalmente emarginada 7
6. Longitud del cuerpo 1.3 a 1.5 mm; coloración castaño oscura; pronoto más ancho en base, lados bastante convexos que convergen al ápice estrechamente redondeado, con mechón de setas ubicado longitudinalmente en la base del pronoto (fig. 24a, e); declive elitral brillante abruptamente redondeado en la base, con carena subserrada en el margen posterolateral, interestrías 2 y 3 con fila de pequeños tubérculos (fig. 24g) Xylosandrus curtulus (Eichhoff)
6’. Longitud del cuerpo 2.1 a 2.9 mm; coloración castaño rojiza; pronoto más ancho en base, lados poco convexos, con setas largas en los bordes y medianas en toda la superficie (fig. 24b, f); declive elitral opaco, muy convexo, densamente cubierto por pequeños gránulos uniforme e irregularmente distribuidos, con carena irregularmente dentada (fig. 24h) Xylosandrus crassiusculus Motschulsky
7. Tercio apical de cara posterior del maza antenal con 2 suturas, cara anterior con segundo segmento más grande, esclerotizado y levemente curvado (fig. 25e, g); cuerpo pequeño, robusto, convexo; longitud del cuerpo 1.8 a 2.1 mm; margen anterior del pronoto armado con 5 dientes medianos Coptoborus villosulus (Blandford)
7’. Tercio apical de cara posterior de maza antenal con 1 o sin suturas marcadas, cara anterior con segundo segmento, si es visible, no esclerotizado y fuertemente curvado (fig. 25f, h); cuerpo alargado, perfil muy levemente convexo a aplanado; longitud del cuerpo 2.0 a 3.2 mm; margen anterior de pronoto sin dientes 8
8. Pronoto subcuadrado (fig. 25b), margen anterior recto, superficie finamente estriada; longitud 2.2 a 2.5 mm; interestrías del declive elitral con filas de tubérculos de diferente tamaño Euwallacea posticus (Eichhoff)
8’. Pronoto más largo que ancho, margen anterior procurvado, superficie lisa; longitud 2.0 3.2 mm; interestrías del declive elitral con tubérculos o espinas 9
9. Declive elitral aplanado y brilloso 10
9’. Declive elitral convexo y opaco o brilloso 11
10. Especie pequeña, cuerpo delgado, longitud del cuerpo 2.0 a 2.9 (fig. 26a); frente con quilla débil ancha que se extiende en zona media, desde epistoma hasta ojos (fig. 26e); declive elitral suavemente empinado, interestría 3 con tubérculo mediano más cerca del ápice que de base del declive (fig. 24g) Xyleborus ferrugineus (Fabricius)
10’. Especie grande, cuerpo robusto, longitud del cuerpo 2.8 a 3.2 (fig. 24b); frente con quilla confinada a zona superior del epistoma (fig. 26f); declive elitral abrupto y empinado, interestría 3 con tubérculo mediano más cerca de base que del ápice del declive (fig. 26h) Xyleborus bispinatus Eichhoff
11. Frente con gránulos pequeños y dispersos (fig. 27e); declive elitral muy empinado y base abrupta (fig. 27c), interestría 1 con tubérculo grande ubicado sobre base protuberante en el tercio basal (fig. 27a, c, g) Xyleborus biconicus Eggers
11’. Frente reticulada sin gránulos; declive eitral con tubérculos pequeños o medianos en las interestrías (fig. 27h) 12
12. Frente fuertemente reticulada y de superficie opaca (fig. 27f); superficie del declive elitral opaca (fig. 27h) Xyleborus affinis Eichhoff
12’. Frente reticulada y de superficie brillante; superficie del declive elitral brillante (fig. 28g, h) 13
13. Frente toscamente reticulada, con línea media elevada y lisa desde el epistoma hasta el nivel superior de los ojos (fig. 28e); declive elitral con interestría 1 con 3 a 5 tubérculos puntiagudos y de tamaño mediano, interestría 2 con 2 o 3 tubérculos puntiagudos y medianos en el cuarto basal y 1 o 2 medianos cerca del ápice, interestría 3 con 3 tubérculos puntiagudos medianos muy espaciados entre sí en el 3 cuarto basal y 3 o más tubérculos pequeños en el cuarto basal (fig. 28g); margen ventrolateral del declive moderadamente elevado y con crenulaciones; coloración del cuerpo castaño rojizo Xyleborus volvulus (Fabricius)
13’. Frente fuertemente reticulada y sin elevaciones (fig. 28f); declive elitral con interestría 1 débilmente elevada, 2 y 3 con fila de tubérculos pequeños y de igual tamaño; coloración del cuerpo castaño rojizo muy oscuro Xyleborus scaber Schedl
Amasa parviseta Knížek et Smith, 2024* (fig. 22a, c, e, g)
Distribución. Argentina: Tucumán; América del Sur: Brasil y Uruguay (Atkinson, 2025; Córdoba et al., 2023, 2025; Flechtmann y Cognato, 2011; Gómez et al., 2017a; Knížek y Smith, 2024; Rainho et al., 2018)
Figura 22. a, Vista dorsal de Amasa parviseta (tomada de Atkinson, 2025); b, vista dorsal de Xyleborinus saxesenii (tomada de Atkinson, 2025); c, vista lateral de A. parviseta (tomada de Atkinson, 2025); d, vista lateral de X. saxesenii (tomada de Atkinson, 2025); e, vista posterior de A. parviseta (tomada de Atkinson, 2025); f, vista posterior de X. saxesenii (tomada de Atkinson, 2025); g, vista ventral de A. parviseta (tomada de Atkinson, 2025).
Comentarios. Es originaria de Australia y se extendió en asociación con especies introducidas de eucalipto. Se informó en Brasil (2015) (Rainho et al., 2018), Uruguay en 2015 (Gómez et al., 2017a), Chile en 2016 (Kirkendall, 2018) y Argentina en 2018 (Córdoba et al., 2023) como A. truncata y A. nr. truncata.
Xyleborinus saxesenii (Ratzeburg, 1837)* (fig. 22b, d, f)
Distribución. Argentina: Buenos Aires, Salta, Santiago del Estero y Tucumán; América del Sur: Brasil, Chile, Ecuador, Paraguay y Uruguay (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Gómez et al., 2017a; Iturre y Darchuck, 1996; Wood, 2007).
Comentarios. Esta es una especie euroasiática y ha sido introducida en todas las regiones templadas y subtropicales del mundo.
Xyleborinus linearicollis (Schedl, 1937) (fig. 23a, c, e, g)
Distribución. Buenos Aires y Tucumán; América del Sur: Brasil (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Xyleborinus sentosus (Eichhoff, 1868) (fig. 23b, d, f, h)
Distribución. Argentina: Tucumán; América del Sur: Brasil, Paraguay y Perú (Atkinson, 2025; Córdoba et al., 2021, 2023, 2025; Smith et al., 2017; Wood, 2007).
Figura 23. a, Vista dorsal de Xyleborinus linearicollis (tomada de Atkinson, 2025); b, vista dorsal de X. sentosus (tomada de Atkinson, 2025); c, vista lateral de X. linearicollis (tomada de Atkinson, 2025); d, vista lateral de X. sentosus (tomada de Atkinson, 2025); e, vista frontal de X. linearicollis (tomada de Atkinson, 2025); f, vista frontal de X. sentosus (tomada de Atkinson, 2025); g, vista posterior de X. linearicollis (tomada de Atkinson, 2025); h, vista posterior de X. sentosus (tomada de Atkinson, 2025).
Xylosandrus curtulus (Eichhoff, 1869) (fig. 24a, c, e, g)
Distribución. Argentina: Salta y Tucumán; América del Sur: Bolivia, Brasil, Colombia, Ecuador, Perú y Venezuela (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Martínez et al., 2019; Smith et al., 2017; Wood, 2007).
Xylosandrus crassiusculus Motschulsky, 1866* (fig. 24b, d, f, h)
Distribución. Argentina: Buenos Aires y Tucumán; América del Sur: Brasil, Guyana Francesa y Uruguay (Atkinson, 2025; Landi et. al., 2017; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Martínez et al., 2019).
Comentarios. Probablemente es originaria de Asia tropical y subtropical, pero se ha extendido ampliamente en regiones cálidas y húmedas de todo el mundo.
Coptoborus villosulus (Blandford, 1898) (fig. 25a, c, e, g)
Distribución. Argentina: Tucumán; América del Sur: Bolivia, Brasil, Colombia, Ecuador, Guyana Francesa, Perú, Trinidad y Tobago y Venezuela (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Smith y Cognato, 2021; Wood, 2007).
Euwallacea posticus (Eichhoff, 1869) (fig. 25b, d, f, h)
Distribución. Misiones y Tucumán; América del Sur: Bolivia, Brasil, Colombia, Ecuador, Guyana, Paraguay, Perú, Surinam, Trinidad y Tobago y Venezuela (Atkinson, 2018; Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Wood, 2007).
Figura 24. a, Vista dorsal de Xylosandrus curtulus (tomada de Atkinson, 2025); b, vista dorsal de X. crassiusculus (tomada de Atkinson, 2025); c, vista lateral de X. curtulus (tomada de Atkinson, 2025); d, vista lateral de X. crassiusculus (tomada de Atkinson, 2025); e, detalle pronoto de X. curtulus; f, detalle pronoto de X. crassiusculus (modificada de Atkinson, 2025); g, vista posterior de X. curtulus (tomada de Atkinson, 2025); h, vista posterior de X. crassiusculus (tomada de Atkinson, 2025).
Figura 25. a, Vista dorsal de Coptoborus villosulus (tomada de Atkinson, 2025); b, vista dorsal de Euwallacea posticus (tomada de Atkinson, 2025); c, vista lateral de C. villosulus (tomada de Atkinson, 2025); d, vista lateral de E. posticus (tomada de Atkinson, 2025); e, vista posterior de la antena de C. villosulus; f, vista posterior de la antena de E. posticus; g, vista anterior de la antena de C. villosulus; h, vista anterior de la antena de E. posticus.
Figura 26. a, Vista dorsal de Xyleborus ferrugineus (tomada de Atkinson, 2025); b, vista dorsal de X. bispinatus (tomada de Atkinson, 2025); c, vista lateral de X. ferrugineus (tomada de Atkinson, 2025); d, vista lateral de X. bispinatus (tomada de Atkinson, 2025); e, vista frontal de X. ferrugineus (tomada de Atkinson, 2025); f, vista frontal de X. bispinatus (tomada de Atkinson, 2025); g, vista posterior de X. ferrugineus (tomada de Atkinson, 2025); h, vista posterior de X. bispinatus (tomada de Atkinson, 2025).
Xyleborus ferrugineus (Fabricius, 1801) (fig. 26a, c, e, g)
Distribución. Argentina: Jujuy, Salta y Tucumán. América del Sur: Bolivia, Brasil, Chile, Colombia, Ecuador, Guyana Francesa, Guyana, Paraguay, Perú, Surinam, Trinidad y Tobago, Uruguay y Venezuela (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Martínez et al., 2019; Smith et al., 2017; Atkinson, 2025).
Xyleborus bispinatus Eichhoff, 1868 (fig. 26b, d, f, h)
Distribución. Argentina: Tucumán; América del Sur: Bolivia, Brasil, Colombia, Ecuador, Guyana Francesa, Paraguay, Perú, Surinam, Trinidad y Tobago y Venezuela (Atkinson, 2025; Córdoba et al., 2021, 2023, 2025; Martínez et al., 2019; Smith et al., 2017).
Xyleborus biconicus Eggers, 1928 (fig. 27a, c, e, g)
Distribución. Argentina: Formosa y Misiones, Tucumán; América del Sur: Bolivia, Brasil, Guyana Francesa y Paraguay (Atkinson, 2025; Córdoba et al., 2023, 2025; Wood, 2007).
Xyleborus affinis Eichhoff, 1868 (fig. 27b, d, f, h)
Distribución. Argentina:Jujuy, Misiones, Salta y Tucumán; América del Sur: Bolivia, Brasil, Chile, Colombia, Ecuador, Guyana Francesa, Guyana, Paraguay, Perú, Surinam, Trinidad y Tobago, Uruguay y Venezuela (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Martínez et al., 2019; Smith et al., 2017).
Xyleborus volvulus (Fabricius, 1775)* (fig. 28a, c, e, g)
Distribución. Argentina:Jujuy yMisiones; América del Sur: Bolivia, Brasil, Colombia, Ecuador, Guyana Francesa, Guyana, Paraguay, Perú, Surinam, Trinidad y Tobago, Uruguay y Venezuela (Atkinson, 2025; Córdoba y Atkinson, 2018; Córdoba et al., 2021, 2023, 2025; Martínez et al., 2019; Smith et al., 2017; Wood, 2007).
Xyleborus scaber Schedl, 1949 (fig. 28b, d, f, h)
Distribución. Argentina: Tucumán; América del Sur: Brasil (Atkinson, 2025; Córdoba et al., 2023, 2025).
Figura 27. a, Vista dorsal de Xyleborus biconicus (tomada de Atkinson, 2025); b, vista dorsal de X. affinis (tomada de Atkinson, 2025); c, vista lateral de X. biconicus (tomada de Atkinson, 2025); d, vista lateral de X. affinis (tomada de Atkinson, 2025); e, vista frontal de X. biconicus (tomada de Atkinson, 2025); f, vista frontal de X. affinis (tomada de Atkinson, 2025); g, vista posterior de X. biconicus (tomada de Atkinson, 2025); h, vista posterior de X. affinis (tomada de Atkinson, 2025).
Discusión
Atkinson (2025) enumera 199 especies de Scolytinae para la Argentina. De estas especies, 56 están presentes en Tucumán, lo cual representa aproximadamente 31% de las especies conocidas de la subfamilia para el país. Consideramos que esta cifra es muy baja en comparación con la fauna de otras regiones neotropicales. El número de especies tucumanas coincide con la abundancia de especies de la tribu Xyleborini documentadas en todo el territorio mexicano, siendo esta tribu uno de los grupos más diversos a nivel mundial (Pérez-Silva et al., 2021).
La diversidad de especies en la provincia de Tucumán, comparada con los países limítrofes ubicados aproximadamente a la misma latitud (Chile, Paraguay y Uruguay) es alta, ya que Atkinson (2025) reporta 52 especies para Chile, 46 para Paraguay y 14 para Uruguay. En Argentina, 10% de las especies son introducidas, mientras que en Uruguay más de la mitad (57%) son especies exóticas. Esto se debe al incremento de terrenos destinados a la plantación de especies forestales arbóreas en Uruguay y al crecimiento del comercio internacional, en particular el intercambio de embalajes de madera, lo que facilita el establecimiento de especies exóticas (Gómez et al, 2017b). En la provincia de Tucumán, 12% son especies introducidas, lo que puede deberse al bajo porcentaje de superficie forestal. La mayoría de estas especies fueron recolectadas en ambientes naturales, son especies establecidas en el país y su origen foráneo está bien documentado (Wood, 1982, 2007).
Figura 28. a, Vista dorsal de Xyleborus volvulus (tomada de Atkinson, 2025); b, vista dorsal de X. scaber (tomada de Atkinson, 2025); c, vista lateral de X. volvulus (tomada de Atkinson, 2025); d, vista lateral de X. scaber (tomada de Atkinson, 2025); e, vista frontal de X. volvulus (tomada de Atkinson, 2025); f, vista frontal de X. scaber (tomada de Atkinson, 2025); g, vista posterior de X. volvulus (tomada de Atkinson, 2025); h, vista posterior de X. scaber (tomada de Atkinson, 2025).
Aunque Wood (2007) realizó una revisión de los Scolytinae para América del Sur, donde se incluyen especies presentes en Argentina, las nuevas especies, los recientes registros y los cambios taxonómicos hacen necesarias claves actualizadas. Además, no se cuenta con claves para la identificación a nivel más local, para la provincia y el país. Uno de los problemas principales de la clave de tribus de Wood (2007) es que intenta seguir sus ideas sobre la filogenia de los grupos y muchas veces ignora características sencillas, empleando algunas más difíciles de interpretar. Como consecuencia, muchos de sus dilemas son demasiado largos y complejos, dificultando su interpretación. Otro problema es que sigue estrictamente un concepto dicotómico sin presentar una matriz de caracteres que esconde similitudes para el reconocimiento. Por ejemplo, en todos los géneros de la tribu Scolytini, la cabeza está descubierta en vista dorsal a pesar de la impresión que presenta la clave en el primer dilema. Hay otros ejemplos que se pueden citar. Aquí hemos intentado usar caracteres menos ambiguos, sin considerar su supuesto valor filogenético.
Las presentes claves taxonómicas, que incorporan los cambios, especies recientemente descritas y nuevos registros, constituyen una importante herramienta para la identificación rápida de las especies de Argentina y América del Sur, lo cual puede ser de gran utilidad en casos de invasión de los cultivos o para futuras investigaciones.
Agradecimientos
A Eduardo Agustín Mendoza (Área Biología Integrativa, Fundación Miguel Lillo) por su valiosa colaboración y asistencia en todas las colectas en el campo. A Pablo Pereyra (Instituto de Iconografía, Área de Zoología, Fundación Miguel Lillo) por su contribución en la toma de fotografías y edición de las mismas.
Referencias
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Propagación in vitro y adaptación de plantas a condiciones ex vitro de Kroenleinia grusonii (Cactaceae)
In vitro propagation and adaptation of plants to ex vitro conditions of Kroenleinia grusonii (Cactaceae)
Dolores Adilene García-González, Juan Pedro Flores-Margez y Pedro Osuna-Ávila *
Universidad Autónoma de Ciudad Juárez, Instituto de Ciencias Biomédicas, Av. Benjamín Franklin Núm. 4650, Zona Pronaf Condominio La Plata, 32310 Ciudad Juárez, Chihuahua, México
*Autor para correspondencia: posuna@uacj.mx (P. Osuna-Ávila)
Recibido: 23 mayo 2025; aceptado: 14 octubre 2025
Resumen
Kroenleinia grusonii (Hildm.) Lodé, es una cactácea endémica del centro de México que enfrenta riesgos como la colecta ilegal o la modificación de su hábitat natural y se encuentra catalogada como en peligro de extinción. El objetivo de este trabajo fue evaluar concentraciones de 6-bencilaminopurina (BAP) sola o combinada con ácido indolacético (AIA) en la formación de brotes de K. grusonii e inducir la formación de raíces para valorar la adaptación de plantas a condiciones ex vitro. La inducción de brotes fue con BAP (0, 1.0, 2.0 y 3.0 mg/L), AIA (0.5 mg/L) y BAP (0, 1.0, 2.0 y 3.0 mg/L) + AIA (0.5 mg/L). Se aplicó AIA 0, 1.0, 2.0 y 3.0 mg/L para enraizamiento y la adaptación ex vitro de las plantas fue con musgo de turba y con una mezcla de musgo de turba y suelo natural (1:1). El mejor tratamiento fue BAP 3.0 mg/L con 5.13 ± 0.9150 brotes por explante. Se obtuvo 89 ± 0.03% de brotes enraizados y la sobrevivencia ex vitro fue de 68.75% en musgo de turba con suelo natural (1:1). Se estableció un protocolo para la propagación clonal de K. grusonii, lo cual representa una contribución para su conservación ex situ.
Palabras clave: Ácido indol-3-acético; 6-bencilaminopurina; Brotes; Cactus; Endémica; Enraizamiento
Abstract
Kroenleinia grusonii (Hildm.) Lodé, is a cactus endemic to central Mexico, that faces risks such as illegal collection or modification of its natural habitat and is listed as endangered. The objective of this work was to evaluate different concentrations of 6-benzylaminopurine (BAP) alone or combined with indoleacetic acid (IAA) in K. grusonii and to induce root formation to assess plant adaptation to ex vitro conditions. Shoot induction was with BAP (0, 1.0, 2.0 and 3.0 mg/L), IAA (0.5 mg/L) and BAP (0, 1.0, 2.0 and 3.0 mg/L) + IAA (0.5 mg/L). AIA 0, 1.0, 2.0 and 3.0 mg/L was applied for rooting and the ex-vitro adaptation of the plants was with peat moss and a mix with peat moss with natural soil (1:1). The best treatment was BAP 3.0 mg/L with 5.13 ± 0.9150 shoots per explant. 89 ± 0.03% of rooted shoots were obtained and the ex-vitro survival was 68.75% in peat moss substrate with natural soil (1:1). A protocol for the clonal propagation of K. grusonii was established, which represents a contribution to its ex-situ conservation.
Keywords: Indole-3-acetic acid; 6-bencylaminoourine; Shoots; Cactus; Endemic; Rooting
Introducción
Kroenleinia grusonii (Hildm.) Lodé (Cactaceae), también conocida como la biznaga barril de oro, es una especie que se utiliza con fines ornamentales, gastronómicos y agropecuarios (Rodríguez et al., 2021). La especie se distribuye principalmente en el centro de México en los estados de Guanajuato, Querétaro, Hidalgo, San Luis Potosí y Zacatecas (Lodé, 2014). Se encuentra en riesgo de desaparecer debido a la destrucción de su hábitat natural, ya que sus poblaciones se encuentran en deterioro por la colecta ilegal, el desarrollo urbano y la construcción de carreteras, entre otros factores (Pérez-Molphe-Balch et al., 2015; Villavicencio-Gutiérrez et al., 2023). Kroenleinia grusonii es un taxón endémico y debido a las adversidades que enfrenta, se encuentra enlistado en la NOM-059-SEMARNAT-2010 bajo la categoría en peligro de extinción (Semarnat, 2019). La UICN (Unión Internacional para Conservación de la Naturaleza) la cataloga como especie en peligro (Guadalupe-Martínez et al., 2013). En consecuencia, es necesario implementar estrategias que permitan la propagación de esta especie para su conservación y protección sin afectar a las poblaciones naturales.
La propagación de las cactáceas se realiza, principalmente, mediante semillas; sin embargo, es necesario buscar otras formas (Manzo et al., 2022), en particular, cuando las especies se encuentran bajo algún estatus de riesgo y no es posible obtener suficientes semillas para iniciar protocolos de propagación (Pérez-Molphe-Balch et al., 2015). Una alternativa es el uso de herramientas biotecnológicas, como el cultivo de tejidos vegetales. Con esta técnica es posible la propagación clonal manteniendo la estabilidad genética clonal y obtener un gran número de plantas con fines de conservación ex situ (Torres-Silva et al., 2021).
Esta herramienta ofrece obtener una rápida multiplicación vegetativa comparado con la propagación tradicional (Mabrouk et al., 2021). Con esta técnica biotecnológica se reduciría el impacto que se ha creado en las cactáceas debido a la sobrecolecta (López-Granero et al., 2021). La técnica se basa en la totipotencia celular que incluye la expresión total del genoma presente en la célula, que permite la multiplicación aumentando el número de individuos (Almeida et al., 2021). Esta cualidad se ve incrementada por la adición de reguladores de crecimiento en el medio de cultivo, especialmente las citoquininas y auxinas, que favorecen el desarrollo de nuevos brotes (Civatti et al., 2017). El potencial regenerativo de las plantas puede variar dependiendo de diversos factores, como el genotipo, el explante utilizado y los reguladores de crecimiento (Torres-Silva et al., 2018).
En cactáceas se ha realizado la propagación in vitro utilizando diferentes explantes como lo son plantas adultas de Opuntia ficus-indica (L.) Mil. (Khalafalla et al., 2007), activación de areolas en Turbinicarpus pseudomacrochele subsp. lausseri (Diers et G. Framk) Glass (de la Rosa-Carrillo et al., 2012), secciones transversales o longitudinales de plántulas de Melocactus glaucescens Buining et Brederoo (Torres-Silva et al., 2018) o brotes como en Echinocactus parryi Engelm. (García-González et al., 2020). Un paso fundamental en el proceso de propagación in vitro es la formación de raíces en los brotes obtenidos (Ivannikov et al., 2022). La inducción de raíces se logra con el uso de auxinas o en ocasiones, sin la necesidad de aplicarlas (Martínez et al., 2016). Por ejemplo, en la cactácea Cereus peruvianus (L.) Mill. los brotes regenerados obtuvieron 100% de enraizamiento en medio de cultivo MS (Murashige y Skoog, 1962) al 25% sin auxinas (Sawsan et al., 2004). Los brotes enraizados son transferidos a condiciones de invernadero, que es la última etapa de la propagación in vitro (Jagiello-Kubiec et al., 2021). El objetivo de la presente investigación fue evaluar diferentes concentraciones de 6-bencilaminopurina (BAP), sola o combinada con ácido indolacético (AIA) en la formación de nuevos brotes de K. grusonii; así como inducir en los brotes la formación de raíces y evaluar 2 sustratos para su adaptación a condiciones ex vitro.
Materiales y métodos
Para la inducción de brotes, se utilizaron plántulas de 90 días de cultivo in vitro de Kroenleinia grusonii como fuente de explantes, que se obtuvieron de experimentos de germinación in vitro de semillas (Osuna-Ávila et al., 2025). A las plántulas se les eliminó la raíz y se utilizaron los ápices completos, los cuales se colocaron en medio de cultivo MS a una concentración de 100% de macro y micronutrientes. El medio de cultivo se suplementó con 30g/L de sacarosa, 7 g/L de agar tipo 1 y el pH se ajustó a 5.7 ± 0.1. Posteriormente, se adicionaron diferentes concentraciones de BAP, AIA y AIA + BAP. Los tratamientos utilizados fueron 8 y consistieron en lo siguiente: control, BAP 1.0, 2.0 y 3.0 mg/L, combinada con o sin AIA. La unidad experimental fue un frasco de 100 ml conteniendo de 20-25 ml del medio de cultivo con 3 repeticiones, se colocaron 5 explantes en cada frasco con 15 explantes por tratamiento y un total de 120 unidades experimentales. Los explantes fueron subcultivados 2 veces en las mismas condiciones cada 60 días. El número total de brotes por cada explante fue registrado después de 120 días de cultivo in vitro.
Los brotes obtenidos de la propagación clonal fueron separados de los explantes utilizando pinzas y se colocaron en medio de cultivo MS al 100% adicionado con diferentes concentraciones de AIA. Los tratamientos fueron 4 y consistieron en lo siguiente: control, AIA 1.0, 2.0 y 3.0 mg/l. La unidad experimental fue el frasco de 100 ml conteniendo 20-25 ml de medio de cultivo con 8 repeticiones, se colocaron aleatoriamente 10 brotes en cada frasco para un total de 80 brotes en cada tratamiento y 320 unidades experimentales. El porcentaje de formación de raíces se evaluó después de 90 días de cultivo in vitro.
Para la adaptación de las nuevas plantas a condiciones ex vitro, se procedió a retirarlas del medio de cultivo utilizando los dedos para ablandar el agar y posteriormente enjuagando las raíces con agua de la llave. A los brotes que formaron raíces, se les midió la longitud de la raíz primaria utilizando un vernier digital y se contaron el número de raíces que se formaron en la base del brote. Se utilizaron 2 sustratos, el primero fue musgo de turba y el segundo fue una mezcla de musgo de turba y suelo natural regosol con textura franco arcillosa, pH de 8 y conductividad eléctrica de 0.45 dS m-1 (IUSS-WRB, 2015) (proporción 1:1). Cada sustrato fue colocado en una charola de 54 cm × 27 cm con 128 cavidades, en donde se colocaron los brotes que formaron raíces. Las plantas se regaron cada 24 h con agua de la llave a capacidad de campo, manteniendo humedad constante del sustrato. Después de 30 días en condiciones ex vitro se evaluó el porcentaje de sobrevivencia.
El análisis de datos fue realizado con el programa estadístico SPSS versión 24.0 (IBM. 2017). Se aplicó la prueba de Kruskal-Wallis no paramétrica para muestras independientes (prueba X²), ya que se trataba de variables discretas (número de brotes y raíces) y de una variable continua (longitud de raíz). La comparación múltiple de promedios se llevó a cabo con la prueba de Bonferroni con un nivel de significancia de 0.05 y los estadísticos descriptivos fueron obtenidos.
Resultados
El número de brotes por explante fue significativamente diferente entre los tratamientos aplicados (p < 0.05). Los mejores resultados de brotación fueron con BAP 3.0 mg/L (fig. 1A), seguido por BAP 1.0 mg/L + AIA 0.5 mg/L (fig. 1B) con un promedio de brotes de 5.13 por explante en cada uno después de 120 días de cultivo in vitro. Cuando el BAP 3.0 mg/L se combinó con AIA 0.5 mg/L, el número de brotes se redujo a 3.6 ± 0.914 (tabla 1). Los resultados no fueron significativos en el grupo control con un promedio de 0.07 ± 0.067 brotes por explante y el AIA 0.5 mg/L con 0.40 ± 0.400 brotes por explante. Los tratamientos de BAP 1.0 mg/L, BAP 2.0 mg/L, AIA 0.5 mg/l + BAP 2.0 mg/L y AIA 0.5 mg/L + BAP 3.0 mg/L, formaron brotes; sin embargo, no fueron significativos con promedios desde 2.8 hasta 3.6 brotes por explante (tabla 1). Los nuevos brotes obtenidos, presentaron diferentes tamaños desde 5 hasta 18 mm de diámetro (fig. 1C).
Tabla 1
Número de brotes obtenidos en la propagación clonal de Kroenleinia grusonii a los 120 días de cultivo in vitro (n = 15).
| BAP mg/L | AIA mg/L | Número de brotes |
| 0 (control) | 0 | 0.07 ± 0.067 b |
| 1.0 | 0 | 3.20 ± 0.890 ab |
| 2.0 | 0 | 3.13 ± 2.109 ab |
| 3.0 | 0 | 5.13 ± 0.9150 a |
| 0 | 0.5 | 0.40 ± 0.400 b |
| 1.0 | 0.5 | 5.13 ± 1.055 a |
| 2.0 | 0.5 | 2.80 ± 1.010 ab |
| 3.0 | 0.5 | 3.60 ± 0.914 ab |
Medias ± el error estándar, letras iguales no son estadísticamente diferentes (n = 80).
Tabla 2
Porcentaje de enraizamiento de brotes, número de raíces por brote y longitud de la raíz principal a los 90 días de cultivo in vitro de Kroenleinia grusonii.
| AIA mg/L | % de enraizamiento | Número de raíces | Longitud de raíces (mm) |
| Control | 66 ± 0.05 b | 3.21 ± 0.385 b | 21.86 ± 1.764 a |
| 1.0 | 77 ± 0.04 ab | 2.89 ± 0.255 b | 19.98 ± 1.631 a |
| 2.0 | 89 ± 0.03 a | 4.68 ± 0.407 a | 22.36 ± 1.583 a |
| 3.0 | 84 ±0.04 a | 3.18 ± 0.293 b | 22.24 ± 1.688 a |
Medias ± el error estándar, letras iguales no son estadísticamente diferentes (n = 80).
La tabla 2 indica que hubo diferencias entre los tratamientos para el porcentaje de inducción de raíces en los brotes cultivados in vitro por 90 días. Para el porcentaje de enraizamiento, se observó que el adicionar AIA al medio de cultivo MS promueve significativamente la formación de raíces en los brotes. Particularmente, con el tratamiento de AIA 2.0 mg/L que resultó ser el mejor con 89 ± 0.03 % de enraizamiento, en comparación con el tratamiento control que mostró 66 ± 0.05 % de brotes enraizados (tabla 2). Por tales motivos, el aplicar AIA 2.0 mg/L resulta ser, aproximadamente, 20% más efectivo para la inducción de raíces que el resto de los tratamientos.
El número de raíces por explante fue significativamente diferente entre tratamientos (p < 0.05). La comparación de promedios muestra que el AIA 2.0 mg/L presentó el mayor promedio de número de raíces con aproximadamente 5 (4.68 ± 0.407) en cada brote. Mientras que el AIA 1.0 mg/L, AIA 2.0 mg/L y el control desarrollaron, en promedio, menos de 3.2 raíces por explante (tabla 2; fig. 2).
En relación con la variable largo de raíz, no se detectó efecto significativo (p = 0.604). El intervalo observado entre los tratamientos fue de 2.38 mm en promedio, es decir, la variable mostró poca variación. Sin embargo, por los resultados obtenidos con AIA 2.0 mg/L, se muestra un promedio ligeramente por arriba que el resto de los tratamientos con 22.36 ± 1.583 mm (tabla 2). Después de 60 días en condiciones ex vitro, el mejor sustrato para promover la aclimatización de las plantas fue el de suelo de turba y suelo natural (1:1). El porcentaje de sobrevivencia que presentó este sustrato fue de 68.75% después de 60 días. Mientras que el sustrato de suelo de turba mostró un porcentaje de sobrevivencia inferior, de 60.93% después de 60 días (fig. 3).
Discusión
La conservación ex situ de cactáceas a través de la propagación clonal es una alternativa viable para las especies que se encuentran en alguna situación de riesgo (Torres-Silva et al., 2021). Con el uso de citoquininas, en algunas cactáceas que no forman brotes laterales, se rompe la dominancia apical y se promueve la formación de nuevos brotes (Lema-Ruminska y Kulus, 2014). Particularmente, con los resultados observados, el uso de BAP 3.0 mg/l promovió la misma cantidad de brotes que la combinación de BAP 1.0 mg/l y AIA 0.5 mg/l, lo cual indica que para obtener resultados significativos no sería necesaria la adición de auxinas al medio de cultivo. Ramírez y Salazar (2016) reportan que la cinetina 10 mg/l es más eficiente en la formación de brotes en los explantes que al combinarla con auxina en Mammillaria petterssonii Hildm. y Coryphanta radians (D. C). Britton et Rose. El resultado es similar en Stenocereus stellatus al utilizar BAP 4.0 mg/L y obtener un promedio de 8 brotes por explante (Martínez et al., 2011). Por el contrario, existen especies que han sido propagadas y que para obtener mayor formación de brotes sí requieren una combinación de citoquinina y auxina. Por ejemplo, en M. geminispina Haw. con cinetina 10 mg/l + AIA 4.0 mg/l obtuvieron 6 brotes por explante (Ramírez y Salazar, 2016). Melocactus glaucescens Buining et Brederoo con BAP 4.0 mg/l + ácido 1-naftalenacético (ANA) 0.25 mg/l presentó 2.5 brotes por explante y sin ANA, un promedio de 0.6 (Torres-Silva et al., 2018). En Echinocactus parryi reportan que es más efectiva la combinación de BAP 2.0 mg/l + AIA 0.5 mg/l al regenerar 2.9 brotes por explante, a diferencia de los tratamientos sin AIA en los que el número de brotes no fue significativo (García-González et al., 2020).
Una consideración importante en Kroenleinia grusonii es que el número de brotes se redujo de 5.13 ± 0.9150 obtenidos con BAP 3 mg/l a 3.6 ± 0.914 al combinar el BAP 3.0 mg/l + AIA 0.5 mg/l. Resultados similares se reportaron para Coryphantha retusa (Pfeiff.) Birtton et Rose, ya que al aplicar BAP 2.0 mg/l el promedio de brotes fue de 8.8 por explante y al combinar BAP 2.0 mg/l + ANA 1.0 mg/l el número de brotes se redujo a 4.2 (Ruvalcaba-Ruíz et al., 2010). Lo mismo sucedió con Pachycereus pringlei (S. Watson) Britton et Rose, con BAP 2.0 mg/l el número de brotes fue de 3.8 ± 0.3 y se redujo a 0.6 ± 0.1 al combinarlo con ANA 0.5 mg/l (Pérez-Molphe-Balch et al., 2002). Los resultados que se obtengan en la propagación clonal dependen del genotipo y de los reguladores de crecimiento utilizados (Montiel-Frausto et al., 2016).
Es de gran importancia establecer un sistema de micropropagación específico para cada especie, no obstante, la respuesta a la formación de raíces y la aclimatización ex vitro puede variar debido a las condiciones de cultivo (Elias et al., 2015). Con los resultados obtenidos en K. grusonii, el tratamiento control mostró una respuesta favorable a la formación de raíces en los brotes. Sin embargo, al utilizar AIA 2.0 mg/l, se logra 30% más de enraizamiento, lo cual significa que se pueden obtener 70 plantas completas. El porcentaje de enraizamiento en K. grusonii se podría considerar alto con la aplicación de AIA debido a que se han reportado porcentajes inferiores de enraizamiento en diferentes especies de cactáceas. Algunos ejemplos son Cereus jamacaru DC. con 70% (Monostori et al., 2012) o Mammillaria vetula subsp. gracilis (Pfeiff.) D. R. Hunt con 20% de enraizamiento (López-Granero et al., 2021). El uso de auxinas como el ácido indolbutírico (AIB) 0.1 mg/l en Hylocereus monacanthus (Hort. ex Lem.) Britton et Rose donde el porcentaje fue mayor que en K. grusonii con 100% de enraizamiento (Montiel-Frausto et al., 2016). En contraste, para diferentes especies y con medios de cultivo sin auxinas, se ha logrado obtener hasta 95% de enraizamiento en C. retusa (Ruvalcaba-Ruíz et al., 2010), Micranthocereus flaviflorus Buining et Brederoo con 98% (Civatti et al., 2017), Mammillaria hernandezii Glass et R. A. Foster con 98.4% de enraizamiento y Mammillaria dixanthocentron Beckeb. ex Mottram con 94% (Lázaro-Castellanos et al., 2018) comparados con los reportados en la presente investigación para K. grusonii.
El uso de AIA 2.0 mg/l es la mejor opción para obtener el mayor número y longitud de raíces en K. grusonii. Resultados similares han sido reportados para H. monacanthus, donde el número de raíces obtenidas en los brotes presentó un promedio de 4.2 por planta, aunque la longitud de la raíz fue aproximadamente 6 veces menor a la de K. grusonii con 3.6 mm (Montiel-Frausto et al., 2016). También Turbinicarpus × mombergeri Říha presentó raíces de menor longitud que las de K. grusonii con 13 mm (Santos-Díaz et al., 2021). En cuanto al porcentaje de sobrevivencia de las plantas aclimatizadas, es importante resaltar que es necesario implementar métodos más eficientes para lograr aumentar el porcentaje de sobrevivencia de las plantas. Ejemplo de ello fue el trabajo realizado por Santos-Díaz et al. (2021), en donde la sobrevivencia de plantas de Turbinicarpus × mombergeri fue de 85% después de 1 año. Otro ejemplo sobresaliente fue lo mostrado por Cortés-Olmos et al. (2023), al reportar 100% de sobrevivencia en plantas de Gymnocalycium cv Fancy. De acuerdo con los resultados obtenidos y con base en literatura consultada, se infiere que las respuestas entre especies pueden llegar a ser muy diferentes, por lo cual se confirma que es necesario establecer protocolos específicos para cada especie y poder aprovechar al máximo el potencial regenerativo de cada una.
Se ha establecido un protocolo de la propagación clonal de Kroenleinia grusonii que inicia con la inducción de brotes, posteriormente el enraizamiento y finaliza con la aclimatización ex vitro. Con los resultados obtenidos en K. grusonii se logró la formación de hasta 5 brotes más con el uso de BAP 3.0 mg/l. Al inducir la formación de raíces en los brotes regenerados, es viable obtener aproximadamente 3 plantas nuevas por cada explante utilizado en la fase de inducción de brotes. Esta información representa un avance para continuar implementando técnicas de propagación in vitro en K. grusonii y que en futuras investigaciones se pueda promover un mayor número de plantas. Los resultados de este estudio, podrían contribuir a los programas de conservación ex situ e in situ de K. grusonii.
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Genetic diversity of Oenothera drummondii (Onagraceae),a dune coastal herb: ecological and evolutionary implications
Diversidad genética de Oenothera drummondii (Onagraceae), una herbácea de dunas costeras: implicaciones ecológicas y evolutivas
Raquel Aurora Hernández-Espinosa a, Jorge González-Astorga a, *,
Yessica Rico b, Juan B. Gallego-Fernández c
a Instituto de Ecología A.C., Red de Biología Evolutiva, Laboratorio de Genética de Poblaciones, Carretera Antigua a Coatepec No. 351, El Haya, 91073 Xalapa, Veracruz, Mexico
b Instituto de Ecología A.C., Centro Regional del Bajío, Red de Diversidad Biológica del Occidente Mexicano, Av. Lázaro Cárdenas No. 253, 61600 Centro, Pátzcuaro, Michoacán, Mexico
c Universidad de Sevilla, Departamento de Biología Vegetal y Ecología, Av. de la Reina Mercedes, 6, 41012 Sevilla, Spain
*Corresponding author: jorge.gonzalez@inecol.mx (J. González-Astorga)
Received: 09 June 2025; accepted: 31 October 2025
Abstract
We studied the diversity and genetic structure of Oenothera drummondii Hook. (Onagraceae), a dune plant with a mixed reproductive system, across 9 populations using 10 microsatellite markers. Plant genetic diversity is governed by intrinsic factors (i.e., reproductive system and dispersal) and extrinsic factors (i.e., population fluctuations and founder effects). We found moderate to low genetic diversity, with southern populations showing lower diversity, and northern populations higher. A separate study on self-compatibility revealed higher selfing in the south and lower in the north, suggesting a latitudinal gradient that may reduce genetic diversity in southern populations. Peripheral populations showed reduced diversity and greater differentiation, likely due to increased isolation and limited gene flow. Central populations near the Texas origin exhibited the highest diversity. Populations from Baja California (O. drummondii subsp. thalassaphila) formed a genetically distinct group, suggesting a separate species. Overall, genetic patterns in O. drummondii genetic diversity reflect historical and ecological influences, including mating system variation, floral traits, and pollinator dynamics. These findings support hypotheses such as center-periphery dynamics and climate-driven historical processes (e.g., post-glacial expansion), which may shape the species’ genetic landscape and suggest possible local adaptations to environmental changes.
Keywords: Neutral genetic diversity; Coastal dunes; Microsatellites; Genetic bottleneck; Latitude; Selfing
Resumen
Evaluamos la diversidad y estructura genética de Oenothera drummondii Hook. (Onagraceae), planta de dunas costeras con sistema reproductivo mixto; se usaron 9 poblaciones y 10 marcadores microsatélites. La diversidad genética en plantas está influenciada por factores intrínsecos (dispersión y sistema reproductivo) y factores extrínsecos (fluctuaciones poblacionales y efecto fundador). Encontramos diversidad genética moderada a baja, con menor diversidad en poblaciones del sur (Ojoshal) y mayor en el norte (Bolívar). Un estudio complementario mostró mayor autocompatibilidad en Ojoshal y menor en Bolívar, sugiriendo un gradiente latitudinal con incremento de autocompatibilidad hacia el sur, con posible reducción de diversidad genética. Las poblaciones periféricas presentaron menor diversidad y mayor diferenciación, asociada a menor flujo génico y mayor aislamiento; las poblaciones centrales, cercanas al origen de la especie en Texas, mostraron mayor diversidad. Las poblaciones de Baja California (O. drummondii subsp. thalassaphila) formaron un grupo genético distinto, que podría representar a una especie separada. En conjunto, los patrones genéticos de O. drummondii reflejan procesos históricos y ecológicos relacionados con el sistema reproductivo, características florales y las dinámicas de polinización, lo cual podría significar adaptaciones locales a cambios ambientales, y patrones de distribución genética asociados con ellos, como la hipótesis de centro-periferia.
Palabras clave: Diversidad genética neutral; Dunas costeras; Microsatélites; Cuellos de botella; Latitud; Autocompatibilidad
Introduction
Genetic diversity may be considered a crucial factor in determining the ability of populations to adapt and evolve, thus increasing their evolutionary potential (He et al., 2024). Understanding the drivers of genetic variation among plant populations is essential in evolutionary biology as this diversity is the foundation for adaptive potential and long-term survival (Chung et al., 2023; Wright, 1969). In plants, the factors that shape genetic variation can be divided into: a) intrinsic biological properties (e.g., the genetic recombination system, including ploidy level, reproductive system, and meiotic behavior; the mode of dispersal and pollination; and the life form), and b) extrinsic dynamic processes (e.g., fluctuations in population size due to bottlenecks, founder effect, invasions, and changes caused by ecological succession) (Duminil et al., 2009). The reproductive system is the intrinsic biological property suggested to be the major driver of genetic diversity in plants (i.e., heterozygosity and population genetic differentiation) (Koelling et al., 2011; Wright, 1969), as it influences the mating patterns within a population by determining the extent to which selfing can occur (Charlesworth, 2006; Raduski et al., 2012). Self-incompatible populations, for instance, are composed of outcrossing plants with a predominantly outcrossing mating system. Conversely, in partially or completely self-compatible plants, the mating system can span from outcrossing to mixed mating to complete selfing (Holsinger, 1991).
Small, low-density populations experiencing habitat fragmentation and isolation, with limited mate or pollinator availability, exhibit increased propensity for self-fertilization (Devaux et al., 2014; Whitehead et al., 2018). In situations where self-fertilization is evolutionarily advantageous, selection will favor changes in traits that facilitate selfing (e.g., reductions in flower size, reduced nectar, pollen and reduced herkogamy) (Opedal, 2018; Shimizu & Tsuchimatsu, 2015; Sicard & Lenhard, 2011). These changes can have consequences at the genetic level, including elevated rates of inbreeding, a reduction in genetic diversity within populations, and increased genetic differentiation among populations (Barrett & Harder, 1996; Ingvarsson, 2002). Furthermore, a principal consequence of predominant self-fertilization is the reduction in heterozygosity and intra-population genetic diversity, coupled with an increase in inter-population differentiation, when compared to self-incompatible plants (Hamrick & Godt, 1996). In contrast, clonal species are expected to exhibit higher levels of heterozygosity than self-incompatible species, but lower levels of polymorphism within the population and higher levels of differentiation between populations (Levin, 2012).
In addition to the intrinsic biological properties of the species, spatial distribution and demography strongly influence processes such as genetic drift, gene flow, and natural selection, which in turn shape the genetic characteristics of populations (Eckert et al., 2008). For example, species range size tends to influence gene flow and genetic differentiation between populations. Large distances between populations with large ranges would be barriers to gene flow (Lawrence & Fraser, 2020). Peripheral populations are predicted to exhibit reduced gene flow, greater isolation, and higher genetic differentiation. In contrast, populations in the center of the distribution would have higher gene flow and be less differentiated (Lawrence & Fraser, 2020). This may be exacerbated if the peripheral populations experience rapid cycles of colonization and extinction, and have associated bottleneck events or founder effects (Eckert et al., 2008). Dispersal ability also influences the amount of gene flow between core and peripheral populations. This has implications for the maintenance of genetic diversity. For example, species with large ranges but limited mobility have reduced gene flow between populations, resulting in greater genetic differentiation (Pelletier & Carstens, 2018).
The genus Oenothera (Onagraceae) is widespread with the majority of species concentrated in western North America (Overson et al., 2023), with some taxa extending to Central Mexico and South America (Wagner et al., 2007). Like other Onagraceae members, Oenothera species originated in the Nearctic region. Their diversification began 20 million years ago during the Miocene, an epoch characterized by colder and drier than the present. These conditions triggered both altitudinal and latitudinal forest retreats, creating ecological opportunities for herbaceous species that prefer open environments and can resist low temperatures and low humidity (Dietrich & Wagner, 1988).
The species of Oenothera contributed significantly to the early development of plant genetics, cytogenetics, and evolutionary biology. Since the work of De Vries in 1900 (Cleland, 1972), a great deal of information has been collected on the ecology, morphology, cytology, and genetics of the genus, giving it a great advantage as a model for study (Greiner & Köhl, 2014), especially on reproductive systems (Johnson et al., 2009; Raven, 1979; Wagner et al., 2007). The intricate evolutionary history of the genus reports several transitions in reproductive systems, which are diverse (Cleland, 1972; Johnson et al., 2011; Rauwolf et al., 2008). In addition, Oenothera has a unique genetic recombination system (Cleland, 1972). In some species, meiosis is accompanied by a rearrangement of chromosome arms that form structures known as rings. These rings segregate into the next generation and have significant implications for inheritance (Golczyk et al., 2014). They act as barriers to homologous recombination and alterations to the linkage balance (Overson et al., 2023; Rauwolf et al., 2008).
Our study focuses on Oenothera drummondii Hook. (Onagraceae), an herbaceous, short-lived perennial species belonging to the subsection Raimmania within Oenothera (Overson et al., 2023). The speciespresents different chromosomal configurations during meiosis, including the formation of bivalents and rings, but is not a permanent translocation heterozygote; so its reproduction is not expected to be clonal (Dietrich & Wagner, 1988). In contrast, the species displays a mixed mating system, encompassing both self- and cross-pollination (Dietrich & Wagner, 1988; Gallego-Fernández & García-Franco, 2021b). Despite exhibiting self-compatibility, its herkogamous large yellow flowers, and elevated stigma suggest predominantly outcrossing, particularly in North American populations (Gregory, 1964). Nevertheless, the species is capable of self-pollination in the absence of pollinators (i.e., sphingid moths and some hymenopterans) or under specific environmental conditions, such as wind, sand burial, and wave action during tropical storms, which are common in dunes and have an impact on the survival of individuals (Dietrich & Wagner, 1988; Gallego-Fernández & García-Franco, 2021; Gregory, 1964). Recent studies in North America showed that levels of self-compatibility vary from population to population, decreasing with increasing latitude (Gallego-Fernández & García-Franco, 2021b). Also, flower size decreased significantly with increasing latitude (Gallego-Fernández & García-Franco, 2021a). Given that all these characteristics could be reflected in the genetic diversity and genetic structure of the populations, we combined existing reports on life history traits, germination and self-compatibility levels, floral traits, dispersal modes, and genetic recombination system with our results of the genetic diversity and structure of 9 populations covering the North American distribution of O. drummondii (Gulf of Mexico and Baja California).
In this study, we addressed the following questions: 1, Does the genetic diversity vary between populations? Because different environmental conditions allow different evolutionary pressures to act in each population, we expect variation in genetic diversity and related genetic parameters such as heterozygosity and inbreeding. 2, Does the predicted variation in genetic diversity relate to intrinsic biological properties of the populations, such as selfing or changes in floral traits? If self-fertilization is the predominant mating system in the population, we expect populations to show smaller flower size, higher inbreeding, lower heterozygosity, and reduced genetic diversity (Ingvarsson, 2002). 3, Do populations closer to the center of the distribution have greater genetic diversity and less genetic differentiation than peripheral populations? Large distances between populations of species with large ranges would be barriers to gene flow (Lawrence & Fraser, 2020). Also, species living in coastal dunes commonly experience rapid cycles of colonization and extinction, and have associated bottleneck events or founder effects (Eckert et al., 2008). This effect would be exacerbated in peripheral populations (Lawrence & Fraser, 2020). We expect that populations at the edge of the distribution would have lower genetic diversity and higher differentiation than central populations.
Materials and methods
Study system
Oenothera drummondii is the sixth member of the series Raimannia within the subsection Raimannia in the genus Oenothera (Dietrich & Wagner, 1988). Its center of origin is in North America (Texas, USA) and consists of 2 subspecies with a disjunct distribution: O. drummondii subsp. drummondii, which is distributed in the coastal dunes of the Gulf of Mexico (from North Carolina in the USA to Campeche in Mexico), and O. drummondii subsp. thalasaphilla along the Pacific coast at the Southern tip of Baja California in Mexico (Dietrich & Wagner, 1988; Hernández-Espinosa et al., 2020). The 2 subspecies can be distinguished morphologically: O. drummondii subsp. drummondii have large flowers and long pubescent leaves, while O. drummondii subsp. thalasaphilla is characterized by having small flowers and small succulent leaves (Dietrich & Wagner, 1988). O. drummondii subsp. drummondii has been accidentally introduced into coastal dune systems around the world, and is considered invasive in several of them (Castillo-Infante et al., 2021; Dietrich & Wagner, 1988). The species is restricted to coastal dunes, inhabiting the dunes from the back beaches to the first dune ridges and embryo-dunes of the Gulf of Mexico and Southeastern USA (from Campeche to North Carolina) (Dietrich & Wagner, 1988; Gallego-Fernández & García-Franco, 2021b; Moreno-Casasola, 1988). From Louisiana to North Carolina in USA, O. drummondii shares its distribution with O. humifusa, another species within subsection Raimmania. Oenothera humifusa is an autogamous, permanent structural heterozygote (PTH), very similar to O. drummondii in growth form and habitat requirements. In the overlapped distribution, 2 basic types can be distinguished, although they are linked by intermediate forms that result from crosses between O. drummondii and O. humifusa (Dietrich & Wagner, 1988; Wagner et al., 2007). To ensure that the collection represented only O. drummondii and excluded the possibility of hybrids, the sampling was made systematically, beginning in Texas, USA, and continuing southward to Tabasco. Some of the populations collected were small and restricted to limited areas, in which all individuals were collected. The most illustrative example was the population of Ojoshal, composed of 9 individuals. We studied 169 individuals from 9 North American populations, and the 2 subspecies of O. drummondii (Fig. 1; Supplementary material: Table S1). Of these, 2 populations are from O. drummondii subsp. thalassaphila (Agua Blanca and Punta Arena del Sur), and the remaining 7 are from subsp. drummondii (Bolívar, South Padre Island, Matagorda, Altamira, La Mancha, Olmeca, and Ojoshal). We had information on self-compatibility and floral traits for 4 of the 9 populations evaluated (Bolívar, South Padre Island, La Mancha, and Ojoshal), since Gallego-Fernández and García-Franco (2021a, b) used the same populations in their studies (Supplementary material: Table S2).
Vegetal material
Leaf material was collected from reproductive adults of each population, for which 5 leaves per plant were sampled in paper bags with silica gel, and the samples were subsequently stored at -20 °C. When population density exceeded 30 individuals, 20 were randomly selected, and if within a population less than 20 individuals were found, as in Ojoshal, the entire population was sampled (Supplementary material: Table S1).
DNA extraction and microsatellite loci amplification
Ten microsatellite loci transferred to O. drummondii from other species of the Oenothera (Hernández-Espinosa et al., 2020) were used in this study: OenhaB105, OenhaD102, OenbidiA_C10, Oenbi2triA_A1, Oenbi2triA_D3, Oenbi39tri10, Oenbi2triA_H1, Oenbi39tri4, Oenbi2triA_E4 and Oenbi39di2. The extraction of genomic DNA was carried out according to 2 protocols: a) CTAB extraction, using 50 mg of dry leaf tissue (González & Vovides, 2002). The extracted DNA was subsequently purified using the PCR Clean-up & Gel Extraction Purification Kit (QIAGEN). b) DNeasy Plant Mini Kit using 20 mg dried leaf tissue following the manufacturer’s instructions (QIAGEN). The quantity and quality of the extracted DNA were verified on 1% agarose gels stained with Red Gel.
Microsatellite amplification was performed using 2 protocols. An initial PCR in a final volume of 10 µL containing total DNA, 2X reaction buffer, 0.2 mM of each dNTP, 1.6 mM MgCl2, 0.5 ng/µL BSA, 0.025 U/µL Taq polymerase, 0.5 µM reverse and forward primers, and 0.5 µM universal primer M13. For sample visualization, each forward primer was modified by the addition of an M13 sequence at the 5’ end (5’-TGT AAA ACG ACG GCC AGT-3’) that is complementary to an M13 primer labeled with either NED (yellow), HEX (green), or 6-FAM (blue) fluorophores. The amplification program consisted of an initial cycle of 3 min at 94 °C; followed by 25 cycles of 94 °C for 40 s, 50 °C for 40 s, and 72 °C for 60 s; 8 cycles of 94 °C for 60 s, 53 °C for 60 s, and 72 °C for 60 s, where hybridization of the M13 primer occurs; and a final extension at 72 °C for 10 min. A second PCR was performed using the Type-it microsatellite kit (QIAGEN), 1 µL of total DNA, and 0.2 µM of the forward, reverse, and M13 primers in cases where no amplification products were obtained from the first PCR after verification in 1% agarose. The amplification program consisted of a 15-minute cycle at 95 °C, followed by 20 cycles at 94 °C for 40 s, 50 °C for 90 s, and 72 °C for 60 s; 15 cycles at 94 °C for 60 s, 53 °C for 60 s, and 72 °C for 60 s. The final extension was performed at 60 °C for 30 min. A standard 400 bp marker and an ABI 3730 Gene Analysis System (Macrogen) were used to analyze the PCR products resulting from amplification. GeneMarker (Softgenetics, State College, PA, USA) was used to identify allele sizes manually. Oenothera drummondii is a diploid species, with amplifications consisting of 1 or 2 alleles per individual.
Statistical analysis
Presence of null alleles was evaluated using the FreeNA software with the EM algorithm (Chapuis & Estoup, 2007). The frequency of null alleles was estimated for each locus and population. FST values were calculated with FreeNA with the ENA algorithm for the correction of null alleles. The presence of clones in the populations was assessed with GenAIEx 6.503 software (Peakall & Smouse, 2012). Deviations from Hardy-Weinberg equilibrium for each locus in each population were assessed by the X2 test using the Benjamin-Hochberg procedure in GenAIEx 6.503 software (Peakall & Smouse, 2012). Linkage disequilibrium (LD) between loci was evaluated in Arlequin 3.5.2 (Excoffier & Lischer, 2010).
The percentage of polymorphic alleles (P), average number of alleles per locus (A), average number of effective alleles (Ae), observed heterozygosity (HO), expected heterozygosity (HE), and fixation index (F) were calculated in GenAIEx 6.503 (Peakall & Smouse, 2012). To estimate the patterns of genetic variation between and within the populations, we calculated the F-statistics (i.e., FIS, FIT and FST) (Wright, 1978). Additionally, FIS per population, considering null allele frequency, was estimated using INEST 2.2 (Chybicki, 2017), applying Bayesian method with 300,000 steps, sampling every 1,000 steps and burn-in of 30,000 steps. For the bottleneck analysis, the Wilcoxon signed rank test was performed for 3 mutation models: Infinite Allele Model (IAM), Stepwise Mutation Model (SMM), and Two-Phase Model (TPM). INEST 2.2 was used to run 100,000 simulations for each mutation. The relationship between latitude and parameters of genetic diversity (A, P, HO and HE) was evaluated using regression models (Sokal & Rohlf, 1995) in the R 4.2.0 program (R Core Team, 2020).
To estimate the molecular variation within and between populations, an analysis of molecular variance (AMOVA) was performed using 1,000 permutations in poppr package in R 4.2.0 software (R Core Team, 2020). To determine isolation by distance (IBD), a Mantel test (Mantel, 1967) was performed between the matrix of genetic differentiation (FST) with INA null allele correction, and population geographic distances (Euclidean) using 999 permutations in the vegan R package (R Core Team, 2020). We measured current dispersal rates (m) using BAYESASS, in which m is interpreted as the fraction of migrants per generation in one population that is derived from another population (Wilson & Rannala, 2003). The Bayesian clustering algorithm implemented in the Structure 2.3.4 software (Pritchard et al., 2010) was used to clarify the genetic structure of the samples by assigning individuals to genetic groups. The simulation was carried out with correlated allele frequencies and with mixture models according to ancestry. All analyses used 100,000 replicates, a Markov chain Monte Carlo (MCMC) burn-in period of 10,000 steps, and 10 replicates per K. The results obtained from Structure were processed with Structure Harvester (Earl & VonHoldt, 2012) to select the optimal K (highest value of ΔK) using the Evanno method (Evanno et al., 2005). Clumpak (Kopelman et al., 2015) and Distruct (Rosenberg, 2004) were used to visualize the bar graphs obtained in Structure. Discriminant analysis of principal components (DAPC) using the R package adegenet with 1,000 permutations for validation was also used to analyze the genetic structure of O. drummondii. It was based on the genetic distances (Cavalli-Sforza & Edwards, 1967) for each pair of populations, using the INA correction (Chapuis & Estoup, 2007) for the presence of null alleles.
Results
Frequency of null alleles
The 10 microsatellite loci amplified successfully in 161 individuals from 9 populations. Statistical tests of Hardy-Weinberg equilibrium for each locus in each population showed that most cases (50 of 78) did not deviate from equilibrium. For each population, the linkage disequilibrium analysis showed significant deviations in at least one locus. Mean null allele frequencies for all populations were low (0.06 ± 0.04). The mean null allele frequencies per population were moderate to low (0.01 ± 0.02-0.13 ± 0.14). Five loci exhibited high allele frequencies (> 0.2) in at least one population when analyzed across all population-locus combinations (Supplementary material: Tables S3, S4).
Genetic diversity
We detected 99 alleles for the 10 microsatellite loci. The total number of alleles detected had an average of 9.9 ± 5.6 alleles per locus, ranging from 4 for OenhaD102 to 23 for OenbidiA_C10. The level of polymorphism was high in all populations (mean 86.67 ± 14.14). The lowest percentage of polymorphism was found in Ojoshal (OJO), where 4 loci were monomorphic. The number of alleles per locus ranged from 1.80 in Ojoshal to 4.60 in Bolívar, with an overall mean number of alleles per locus of 2.28 ± 0.48. We found a total of 35 private alleles, ranging from 1 in La Mancha to 8 in Bolívar and Agua Blanca. The average effective number of alleles was 2.28 ± 0.48. It ranged from 1.58 in Olmeca to 3.06 in Bolívar (Table 1).
Expected heterozygosity ranged from 0.283 in Olmeca to 0.599 in Matagorda. The overall mean was 0.442 ± 0.104. The observed heterozygosity was also lowest in Olmeca (0.221), and highest in Matagorda (0.547). The mean total observed heterozygosity was 0.409 ± 0.097. The average observed heterozygosity was lower (HO < HE) than expected, indicating an overall homozygote excess. This is consistent with the results of the Hardy-Weinberg equilibrium test, which showed heterozygote deficiency for at least 1 locus in all populations. However, the Hardy-Weinberg equilibrium test also showed heterozygote excess in 4 populations: South Padre Island, Olmeca, Ojoshal, and Punta Arena del Sur. Of these, Ojoshal showed heterozygote excess at 5 of the 6 polymorphic loci (Table 1).
Table 1
Parameters of neutral genetic diversity of O. drummondii at 10 microsatellite loci. N = Sample size; P = percentage of polymorphic loci; A = mean allele number by locus; Ae = mean effective allele number; Ap = private alleles; HO and HE = mean observed heterozygosity, and mean expected heterozygosity; F = fixation index. Hardy-Weinberg equilibrium deviations (p < 0.05), test is the number of polymorphic loci evaluated, and (-) is the number of loci with deficiency and excess of heterozygotes (+).
| Population | N | P | A | Ae | Ap | Ho | HE | F | H-W deviations | ||
| Test | (+) | (-) | |||||||||
| Bolívar | 20 | 100 | 4.60 | 3.06 | 8 | 0.410 | 0.557 | 0.264 | 10 | – | 4 |
| Matagorda | 19 | 100 | 4.20 | 2.75 | 3 | 0.547 | 0.599 | 0.086 | 10 | – | 1 |
| South Padre Island | 20 | 80 | 3.60 | 2.54 | 2 | 0.385 | 0.433 | 0.110 | 8 | 1 | 1 |
| Altamira | 20 | 100 | 3.10 | 2.27 | 1 | 0.485 | 0.487 | 0.003 | 10 | – | 2 |
| La Mancha | 20 | 90 | 3.10 | 2.28 | 3 | 0.420 | 0.464 | 0.094 | 9 | – | 3 |
| Olmeca | 19 | 70 | 3.00 | 1.58 | 6 | 0.221 | 0.283 | 0.219 | 7 | 1 | 2 |
| Ojoshal | 9 | 60 | 1.80 | 1.61 | 0 | 0.489 | 0.301 | -0.623 | 6 | 5 | 1 |
| Agua Blanca | 17 | 90 | 4.10 | 2.31 | 8 | 0.324 | 0.435 | 0.256 | 9 | – | 4 |
| Punta Arena del Sur | 18 | 90 | 3.30 | 2.11 | 4 | 0.400 | 0.420 | 0.047 | 9 | 1 | 2 |
| Mean ± SD | 18 ± 3.54 | 86.7 ± 14.1 | 3.42 ± 0.83 | 2.28 ± 0.48 | 0.41 ± 0.097 | 0.44 ± 0.10 | 0.05 ± 0.27 |
Table 2
Bottleneck estimation for 9 populations of O. drummondii. Significant values are shown in bold. SSM: Step mutation model, TPM: two-phase model.
| Population | FIS (INest) | FIS (INest) 95% HDPI | Bottleneck test (p-value) | |
| SSM | TPM | |||
| Bolívar (BOL) | 0.2061 | 0.2061-0.3218 | 0.385 | 0.3473 |
| Matagorda (MAT) | 0.1221 | 0.0023-0.2274 | 0.0656 | 0.0421 |
| South Padre Island (SPA) | 0.0356 | 0.0011-0.0715 | 0.1913 | 0.1907 |
| Altamira (ALT) | 0.0051 | 0.0000-0.0049 | 0.0322 | 0.0245 |
| La Mancha (MAN) | 0.0214 | 0.0044-0.0568 | 0.0820 | 0.0486 |
| Olmeca (OLM) | 0.2446 | 0.2277-0.3479 | 0.9845 | 0.9845 |
| Ojoshal (OJO) | 0.0133 | 0.0000-0.0217 | 0.0264 | 0.0268 |
| Agua Blanca (AGB) | 0.0977 | 0.0586-0.1955 | 0.973 | 0.9625 |
| Punta Arena del Sur (PAS) | 0.0196 | 0.0000-0.471 | 0.5443 | 0.4556 |
The coefficients of inbreeding obtained using the INEST ranged from 0.0051 in Altamira to 0.2446 in Olmeca (Table 2). Three models were evaluated to detect the presence of recent genetic bottlenecks. The infinite allele model (IAM), the two-phase model (TPM), and the step mutation model (SSM). The SMM model has been suggested as a statistically conservative approach to detect microsatellite bottlenecks (Luikart et al., 1998). However, given the instability and wide range of mutation rates inherent to these markers, the TPM model may be the one that best explains the predominant mutation process at most loci (Di Rienzo et al., 1994). According to the TPM model, 4 populations in O. drummondii show possible recent bottlenecks: Altamira, Ojoshal, Matagorda and La Mancha (Table 2). Regression analyses revealed significant latitudinal clines in genetic diversity, with both alleles per locus (r = 0.69, p < 0.005) and expected heterozygosity (r = 0.65, p < 0.005) showing strong negative relationships with latitude. Latitude significantly predicted the percentage of polymorphic loci (r = 0.4, p < 0.05), whereas no significant relationship was observed for heterozygosity (Supplementary material: Table S5).
Genetic structure
When evaluating the distribution of genetic variation within and among populations, mean values of global inbreeding (FIT = 0.416 ± 0.163) and local inbreeding (FIS = 0.059 ± 0.195) indicated significant heterozygosity deficits. The mean value of genetic differentiation (FST) was 0.376 ± 0.132, indicating that 38% of the genetic variation in O. drummondii was due to differences between populations. Furthermore, genetic differentiation between pairs of populations was mostly high (> 0.15). Matagorda and Bolívar had the lowest differentiation (0.06), followed by Punta Arena del Sur and Agua Blanca (0.17), and Bolívar and South Padre Island (0.18). The analysis of molecular variance (AMOVA) showed that most genetic variation was found among individuals (54%), followed by that found among populations (38%) (Table 3).
The Mantel test detected significant isolation by distance (r = 0.64, p = 0.002). The BAYESASS analysis revealed high diagonal m-values (G1 = 0.972 – G3 = 0.980), suggesting high gene flow within the individual groups. Higher gene flow was observed from G1 to G2 (m1 to 2 = 0.009) (Fig. 2). The results of the Bayesian clustering are shown in Supplementary material: Fig. S1. The most probable K with the highest value of ΔK = 479.023 was K = 3, followed by K = 4 with ΔK = 101.129. For K = 3, 3 clearly defined genetic groups were observed, with a very low proportion of admixture among them. In this case, the first group consisted of the Texas populations: Bolívar, Matagorda, and South Padre Island, and the northernmost Mexican population, Altamira. The second group included the 3 remaining Mexican populations in the Gulf of Mexico: La Mancha, Olmeca and Ojoshal, while the third group comprised the 2 populations on the Baja California Peninsula. For K = 4, the clustering is similar to K = 3. The main difference in this case was that the fourth cluster split the previous cluster containing the Texas and Altamira populations, indicating greater genetic correspondence between Bolívar and Matagorda (populations further north in the Gulf) and between South Padre Island and Altamira. Admixture was low, indicating that Oenothera populations are highly structured.
The Discriminant Analysis of Principal Components (DAPC), like the Bayesian method, also showed 4 groups (Fig. 1C; Supplementary material: Fig. S2). Consistent with the results of Structure, one group was formed by the populations of Baja California (Punta Arena del Sur and Agua Blanca). However, in this case, La Mancha was no longer grouped together with Ojoshal and Olmeca, which formed the second group. Altamira and La Mancha formed a third group that is closer to the fourth formed by the Texan populations (Bolívar, Matagorda, and South Padre Island). The DAPC results indicated the possibility that the Texas group was the ancestral group of the central Gulf of Mexico populations (Altamira and La Mancha). The other 2 groups were markedly more differentiated.
Table 3
Molecular variance analysis (AMOVA) of Oenothera drummondii populations.
| Source of variation | df | Sum sq | Mean sq | Variance component | Total variance (%) | p values |
| Between populations | 8 | 841.68 | 105.21 | 2.78 | 37.39 | 0.001 |
| Among populations | 153 | 807.84 | 5.28 | 0.61 | 8.21 | 0.001 |
| Within individuals | 162 | 657 | 4.05 | 4.05 | 54.39 | 0.001 |
| Total | 323 | 2,306.53 | 7.14 | 7.45 | 100 |
Discussion
In this research, we evaluated the genetic variation between and within populations of Oenothera drummondii throughout its range in North America, for which we used microsatellite markers. These markers demonstrated sufficient variability to resolve population genetic structure at high resolution, consistent with a previous study (Hernández-Espinosa et al., 2020). We also found that all populations were in linkage disequilibrium for at least one locus. Deviations may result from limited homologous recombination during meiosis when chromosomal rings are formed (Rauwolf et al., 2008). In O. drummondii, as in many Oenothera species, this limited recombination is reported to cause changes in the linkage equilibrium (Cleland, 1972; Rauwolf et al., 2008; Raven, 1979), as described in other ring-forming species of the genus: O. biennis (Larson et al., 2008; Levin, 1975; Levy & Levin, 1975), O. harringtonii (Skogen et al., 2012), O. hartwegii and O. gayleana (Lewis et al., 2016). The unique recombination system in the Oenothera is an attribute that has allowed them to colonize new environments from their origin in the Miocene (ca. 20 Ma), until recent times, successfully establishing themselves first in the Neotropics and then in Europe (Dietrich et al., 1997; Wagner et al., 2007).
Compared to other plants with similar characteristics in lifeform, reproductive system, geographic range, and dispersal mechanism, we found that Oenothera drummondii showed moderate to low genetic diversity. This is a short-lived perennial, self-compatible species with a mixed reproductive system and wide distribution (Dietrich & Wagner, 1988), and a variety of dispersal mechanisms (Gallego-Fernández et al., 2021). Mean levels of genetic diversity reported for other plants with mixed reproduction (N = 15, mean HE = 0.60), short-lived perennial life form (N = 29, mean HE = 0.55), wide distribution (N = 31, mean HE = 0.62), and water or wind dispersal (N = 28, mean HE = 0.61) (Nybom & Bartish, 2000), were higher than the level found in O. drummondii (mean HE = 0.442).
Historically, genetic diversity in Oenothera has been attributed to structural genomic rearrangements, particularly hybridization and reciprocal translocations that alter linkage relationships translocations (Cleland, 1972; Rauwolf et al., 2011), as the genomes in Oenothera were considered to be essentially “non-recombining” due to the formation of chromosomal rings during meiosis (Cleland, 1972; Golczyk et al., 2014; Rauwolf et al., 2008). In contrast, the reported genetic diversity in Oenothera species appears strongly influenced by their predominant reproductive system. O. biennis, a permanent translocation heterozygote (PTH), maintains higher genetic variation than O. drummondii, which exhibits a mixed mating system (mean HE = 0.69 vs. 0.44) (Larson et al., 2008). This agreed with the expectation of the “functional” asexual reproductive system of O. biennis, in which autogamy is almost complete and offspring are mostly clonal (Mather, 1943; Stebbins, 1957). Consistent with expectations for obligate sexual reproduction, self-incompatible, bivalent-forming Oenothera species (O. harringtonii with mean HE = 0.77, O. gayleana with mean HE = 0.53 and O. hartwegii with mean HE = 0.49 vs O. drummondii with HE = 0.44) maintained significantly higher genetic diversity than the mixed-mating O. drummondii (HE = 0.44) (Levin, 1975; Rhodes et al., 2014; Skogen et al., 2019).
In terms of its reproductive system, the self-compatible O. drummondii generally has large flowers and an elevated stigma, suggesting a high propensity for sexual reproduction, although it can self-pollinate in the absence of pollinators or in situations where self-fertilization is evolutionarily advantageous (Gregory, 1964; Sicard & Lenhard, 2011). Gallego-Fernández and García-Franco (2021b) assessed the level of self-compatibility of O. drummondii in the Gulf of Mexico, through germination essays. They evaluated 4 of the 9 populations in this study (Bolívar, South Padre Island, La Mancha and Ojoshal), and found that self-compatibility increased in low-latitude populations (La Mancha and Ojoshal), while lower for higher-latitude populations (Bolívar and South Padre Island). These results suggest a latitudinal shift towards greater selfing to the south of the distribution. It has been suggested that greater selfing in populations leads to a decrease in genetic diversity (Ingvarsson, 2002), which is consistent with the pattern of genetic variation found in this study. In agreement with the results of Gallego-Fernández and García-Franco (2021a), we found that populations with higher reported self-compatibility also had lower genetic diversity. On the other hand, in partial selfing species, reductions in genetic diversity are associated with changes in traits that facilitate selfing (i.e., reduced flower size, reduced nectar, pollen and reduced herkogamy) (Barrett & Harder, 1996; Shimizu & Tsuchimatsu, 2015; Tedder et al., 2015). According to the theory, it would be expected that the populations of O. drummondii with a higher self-compatibility, and lower genetic diversity (i.e., Ojoshal) would also show a tendency to reduce flower size, and populations with lower self-compatibility, and higher genetic diversity (i.e., Bolívar) would show an increase in flower size. However, Gallego-Fernández and García-Franco (2021a) found the opposite: populations showed larger flower sizes following a latitudinal gradient from subtropical (i.e., Bolívar and South Padre Island) to tropical (i.e., La Mancha and Ojoshal) climates. This result suggests that in O. drummondii there is no latitudinal correlation between a reduction in flower size and an increase in selfing, and thus to a reduction in genetic diversity. Differences in floral traits among O. drummondii populations appear to respond to both biotic (i.e., pollinators) and abiotic factors (i.e., temperature, precipitation), suggesting local adaptations to environmental changes (Gallego-Fernández & García-Franco, 2021a).
In O. drummondii the genetic diversity showed a significant increase with latitude, with its lowest values in Ojoshal and Olmeca, while the higher values were found in Bolívar and Matagorda, populations in the highest latitudes. The genus Oenothera appears to have originated in northern Mexico and the adjacent USA, and subsection Raimmania, to which O. drummondii belongs, is centered in this area (Dietrich & Wagner, 1988). Since older clades tend to have more genetic diversity (Willi et al., 2018), it would be expected that the greatest genetic diversity would be found in populations at latitudes closest to the center of origin (i.e., Bolívar and Matagorda) and that it would lower as populations moved away to the periphery of the distribution, as was found here. In addition to reducing population genetic variation through genetic drift, habitat suitability often decreases from the core to the edge of a species’ geographic range, leaving edge populations small and isolated (Sagarin & Gaines, 2002), as in Ojoshal. However, despite the small population size of Ojoshal, inbreeding was low, presumably because of its high selfing index (Gallego-Fernández & García-Franco, 2021b). Inbreeding depression is predicted to be lower in selfing species because recessive deleterious alleles are expected to be efficiently “purged” from populations (Winn et al., 2011), and selfing favors the selection of new recessive beneficial mutations (Burgarella & Glémin, 2017; Charlesworth et al., 1993).
The pattern of distribution of the genetic diversity found in Oenothera drummondii seems to be shaped by demographic and evolutionary processes (Durka, 1999; Hewitt, 1996). Historical climate-driven changes are known to still affect the present-day genetic diversity (Alvarez et al., 2009; Hewitt, 2000). As a result of sequential founder events during post-glacial recolonization, contemporary populations show reduced genetic diversity and increased genetic differentiation, especially at range margins (Eckert et al., 2008; Sagarin & Gaines, 2002). The above pattern was found in 512 North American herbs, showing that the shifting and fragmentation of species’ geographical ranges in the past 20,000 years has played an important role in shaping the genetic variation of contemporary populations (Benavides et al., 2019).
Regarding the populations in Baja California (i.e., Agua Blanca and Punta Arena del Sur), the results of both genetic diversity and genetic structure showed that they formed an independent genetic group, as expected given that the populations belong to a different subsp. of O. drummondii. However, based on the genetic differences detected and the evidence of differences in life history, vegetative and floral characteristics observed in the field at the time of collection, it is suggested that O. drummondii subsp. thalassaphila is a distinct species from O. drummondii subsp. drummondii (Table 4). A detailed taxonomic study and additional genetic data from new populations of O. drummondii subsp. thalassaphila will be needed to support this.
Table 4
Distinctive morphological attributes between O.drummondii subsp. drummondii and O. drummondii subsp. thalassaphila. Data taken from Benavides et al. (2019).
| O.drummondii subsp. drummondii | O. drummondii subsp. thalassaphila | |
| Habit | Annual herb | Perennial sudshrub |
| Stem pubescence | Strigillose to villous | Strigillose |
| Glandular puberulent hairs | Present | Absent |
| Cauline leaves | 1-8cm, oblanceolate to obovate, densely villous, margins entire to remotely sinuate dentate | 1-4.5cm, oblong lanceolate to oblanceolate, strigillose, marginis entire to coarsely dentate |
| Floral tube | 2.5-5cm, strigillose to densely villous | 2-3.5cm, strigillose |
| Sepals | 2-3cm, strigillose to villous | 1.3-2.5cm, strigillose, occasionally red doted |
| Petal length | 2.5-4.5cm | 2-3.5cm |
| Capsule | 2.5-5.5cm, strigillose to villous | 2-4cm, strigillose |
| Seeds | 1.1-1.7mm | 1.5-2mm, smooth |
We conclude that the genetic diversity and genetic structure across the range of O. drummondii in North America suggest a latitudinal shift towards increased selfing and a decline in genetic diversity in the low-latitude populations. However, there is no positive relation between reductions in floral traits and increased selfing in the populations. Differences in floral traits appear to respond to biotic and abiotic factors, suggesting local adaptations to environmental changes. In O. drummondii, genetic parameters may reflect historical processes like climate-driven changes (i.e., the last glaciation), and patterns of genetic distribution associated with them, such as the center-peripheral hypotheses. The genetic attributes evaluated here suggest that populations of O. drummondii subsp. thalassaphila may represent a distinct species, but further evaluation will be required to confirm this hypothesis.
Acknowledgments
This study was supported by the Consejo Nacional de Ciencia y Tecnología (Conacyt) by a scholarship (2019-000037-02NACF-29365), as part of a doctoral thesis directed by Jorge González-Astorga. This study was also supported by the Ministerio de Economía y Competitividad, Spain (MINECO Project CGL2015-65058-R co-funded by FEDER). We sincerely thank the assistance in fieldwork to collect plant materials: Rusty Feagin, José García-Franco, Alejandro Espinosa de los Monteros and Anwar Medina-Villareal. We also thank Janet Nolasco-Soto for the assistance in laboratory methods, and two anonymous reviewers for their comments and observations on the manuscript.
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Ensambles de polen moderno y fósil en la sierra Occidental de Jalisco; ~ 1,500 años de historia de la vegetación y cambio ambiental
Modern and fossil pollen assemblages in the Sierra Occidental of Jalisco; ~ 1,500 years of history of the vegetation and environmental change
Ana Patricia del Castillo-Batistaa, *, Blanca Lorena Figueroa-Rangela, Socorro Lozano-Garcíab, Ramón Cuevas-Guzmána, Miguel Olvera-Vargasa y Lia Hueso-Vidrioc
a Universidad de Guadalajara, Laboratorio de Paleoecología y Cambio Climático, Departamento de Ecología y Recursos Naturales, Centro Universitario de la Costa Sur, Av. Independencia Nacional Núm. 151, 48900 Autlán de Navarro, Jalisco, México
b Universidad Nacional Autónoma de México, Instituto de Geología, Laboratorio de Paleoecología, Paleoclimatología y Cambio Climático, Departamento de Dinámica Terrestre Superficial, Ciudad Universitaria, Coyoacán, 02376 Ciudad de México, México
c Universidad de Guadalajara, Ingeniería en Recursos Naturales y Agropecuarios, Departamento de Ecología y Recursos Naturales, Centro Universitario de la Costa Sur, Av. Independencia Nacional Núm. 151, 48900 Autlán de Navarro, Jalisco, México
*Autor para correspondencia: ana.delcastillo@academicos.udg.mx (A.P. del Castillo-Batista)
Recibido: 15 agosto 2025; aceptado: 30 octubre 2025
Resumen
La sierra de Cacoma, Jalisco, es una región de alta diversidad florística y endemismo situada en una zona de transición biogeográfica (2,119 m snm), donde convergen elementos templados y tropicales. Analizamos la relación entre la composición taxonómica del polen moderno y fósil, así como los cambios ambientales en un bosque de Pinus–Quercus–Abies de La Cumbre de Guadalupe (Talpa de Allende, Jalisco, México). Se analizaron 22 muestras de polen moderno, 62 de polen fósil y la estructura de la vegetación en 14 parcelas de 500 m2. Aplicamos los índices de asociación palinológica de Davis, la diversidad de Hill y una similitud basada en la distancia cordal. Los últimos ~ 1,580 años muestran un intervalo de mayor similitud (~ 850 y 400 años cal AP) intercalado por episodios de cambio (~ 1,500, 1,000 y 300 años cal AP), congruentes con transiciones climáticas del Holoceno tardío. En el periodo reciente, incendios y deforestación han favorecido la homogeneización de la vegetación. Estos resultados documentan las respuestas de bosques templados a forzamientos climáticos y antrópicos a escala centenaria y resaltan el valor de integrar lluvia de polen moderno, registros fósiles y métricas de diversidad.
Palabras clave: Actividad antrópica; Bosques templados; Holoceno tardío; Lluvia de polen
Abstract
The Sierra de Cacoma, Jalisco is a region of high floristic diversity and endemism located in a biogeographic transition zone (2,119 m asl), where temperate and tropical elements converge. We analyzed the relationship between the taxonomic composition of modern and fossil pollen and environmental changes in a Pinus-Quercus-Abies forest at La Cumbre de Guadalupe (Talpa de Allende, Jalisco, Mexico). We analyzed 22 modern pollen samples, 62 fossil pollen samples, and vegetation structure in 14 plots (500 m² each). We applied Davis palynological association indices, Hill diversity, and a dissimilarity analysis based on chord distance. The past ~ 1,580 years show an interval of greater similarity (~ 850-400 cal yr BP) interspersed with episodes of change (~ 1,500, 1,000, and 300 cal yr BP), consistent with Late Holocene climatic transitions. In the recent period, fires and deforestation have promoted vegetation homogenization. Overall, these results document the response of temperate forests to climatic and anthropogenic forcing at centennial scales and highlight the value of integrating modern pollen rain, fossil records, and diversity metrics to interpret vegetation and environmental history.
Keywords: Anthropogenic activity; Temperate forests; Late Holocene; Pollen rain
Introducción
Las comunidades vegetales actuales incorporan señales del cambio climático acumuladas a múltiples escalas temporales que son evidentes en la distribución, fenología, dinámica demográfica y composición florística, pero su estructura y trayectoria también responden a otros motores de cambio, como el uso de suelo, régimen de incendios, especies invasoras, perturbaciones antrópicas y variabilidad natural (Jackson y Overpeck, 2000). Estos forzamientos configuran patrones contemporáneos, en los que el clima es determinante. En este contexto, el análisis de polen fósil preservado en sedimentos de hondonadas forestales es una herramienta robusta para reconstruir la historia de la vegetación y su ambiente (Calcote, 1995; Figueroa-Rangel et al., 2020). Los registros polínicos documentan cambios en la composición taxonómica a través del tiempo (Figueroa-Rangel et al., 2016) y mediante un enfoque espacio-temporal, la paleoecología permite interpretar la dinámica de los ecosistemas y las respuestas de la vegetación frente a variaciones y perturbaciones ambientales (Trivi de Mandri et al., 2006).
La reconstrucción paleoecológica se basa en el supuesto de que existe una estrecha relación entre las especies de plantas presentes en el bosque y el polen que éste produce (Erdtman, 1943; Faegri e Iversen, 1989). El polen se libera y se dispersa durante la maduración y apertura de los sacos polínicos y es depositado como sedimento en capas del suelo donde se preserva en condiciones ambientales adecuadas (húmedas y anóxicas). Al analizar los conjuntos de polen fósil presentes en estos sedimentos, es posible inferir la composición de la vegetación que existió en el pasado. Así mismo, la lluvia de polen moderno desempeña un papel importante para comprender la representación del polen en las secuencias sedimentarias y calibrar la interpretación basada en estos registros. El análisis de los conjuntos de polen moderno permite establecer la relación entre la abundancia relativa del polen y la presencia de diferentes especies de plantas en la vegetación actual (Rodríguez-Pérez et al., 2025). Lo que contribuye a mejorar la interpretación de los conjuntos de polen fósil y una comprensión más precisa de los cambios de la vegetación a lo largo del tiempo (Lozano-García et al., 2014).
La abundancia de polen en los sedimentos resulta de la interacción de factores biológicos de las especies fuente como la productividad, modo de dispersión y preservación a través del tiempo (Davis, 1984), además de factores ecológicos locales que integran la estructura de la vegetación desde el rodal hasta su entorno inmediato (diámetro y altura de los árboles, área basal, densidad, frecuencia, composición y fenología), modulando el transporte y la deposición del polen (Davies y Fall, 2001; Faegri e Iversen, 1989), y así mismo de las condiciones físicas de los sedimentos del sitio como el tamaño y morfología del sustrato, oxidación y frecuencia de inundación (Davies y Fall, 2001; Faegri y van der Pijl, 1979; Proctor et al., 1996); finalmente, de las características climáticas y la calidad de la identificación palinológica (Escarraga-Paredes et al., 2014). Por lo tanto, el análisis del polen fósil y el estudio de las lluvias de polen moderno son herramientas complementarias que permiten reconstruir la historia de la vegetación y su ambiente, especialmente en regiones donde no se dispone de información sobre la producción y dispersión de polen (Wright, 1967).
Para comprender mejor la dinámica de la vegetación, es necesario abordarla desde una perspectiva espacio-temporal, donde diferentes procesos físicos y biológicos influyen en los patrones observados en cada escala temporal. Las respuestas de la vegetación a corto plazo se pueden interpretar a través de eventos como incendios, huracanes, enfermedades y extracción de madera, mientras que a largo plazo están involucrados los mecanismos de perturbación ambiental y el desarrollo evolutivo (Delcourt et al., 1983). La escala espacial permite comprender cómo los diferentes tipos de vegetación cercanos geográficamente responden de forma diferenciada en eventos de corto o largo plazo, tales como los regímenes de perturbación y fluctuaciones climáticas (Figueroa-Rangel y Olvera-Vargas, 2018). En este sentido, se han realizado diversos estudios con el uso de polen fósil para reconstruir la historia de la vegetación y el clima en el occidente de México, durante el Holoceno tardío (Figueroa-Rangel et al., 2008; Lozano-García et al., 2021). Estos estudios han proporcionado herramientas para evaluar la paleovegetación y los escenarios climáticos en las regiones de alta montaña y en las regiones tropicales. Sin embargo, se ha prestado poca atención a la comprensión de las relaciones cuantitativas entre la lluvia de polen y la vegetación moderna para la interpretación de la historia de la vegetación y el clima con base en datos de polen fósil.
Un área de gran importancia para realizar estudios que combinen registros de polen fósil y moderno es la región de la sierra de Cacoma, ubicada en el occidente de México. Esta región se destaca por su alta diversidad biológica y endemismo de flora y fauna (Conabio, 2010). Se encuentra en una zona de transición biogeográfica con un rango altitudinal de 650 a 2,740 m, con zonas de transición entre tipos de vegetación templada y tropical (INEGI, 2015). Las zonas de mayor elevación comprenden topografías abruptas, con valles y barrancas profundas que forman microclimas que han servido de refugio para varias comunidades de plantas (Amador-Cruz et al., 2024). Por lo tanto, el presente estudio tiene como objetivo evaluar, mediante el análisis de ensambles de polen moderno y su correspondencia con la estructura de la vegetación, si el espectro de polen fósil refleja los patrones de la distribución actual de la vegetación. Para ello, se plantean los siguientes objetivos: 1) caracterizar la composición taxonómica y la representatividad de la lluvia de polen moderno en el bosque de Pinus–Quercus–Abies en la localidad de La Cumbre de Guadalupe en el municipio de Talpa de Allende, Jalisco, identificando los principales taxones y su relación con la vegetación fuente, 2) analizar la correspondencia entre la estructura de la vegetación actual y la señal polínica mediante análisis cuantitativos de la abundancia del polen moderno y 3) comparar la similitud taxonómica entre los ensambles de polen moderno y fósil de los últimos ~ 1,580 años. Este enfoque permitirá fortalecer la interpretación de los registros paleoecológicos en la región y contribuir al conocimiento sobre las relaciones entre la vegetación actual y la señal polínica moderna en los ecosistemas de alta montaña.
Materiales y métodos
El área de estudio se localiza en la sierra de Cacoma al occidente de Jalisco, México, en la localidad denominada La Cumbre de Guadalupe, una región de confluencia entre la cordillera del Eje Neovolcánico Transversal y la sierra Madre del Sur (fig. 1). Corresponde a una elevación de 2,119 m y se extiende hasta el litoral del Pacífico y la costa del occidente de México (INEGI, 2015). Es una zona de gran complejidad geomorfológica y litológica con afloramientos rocosos formados a partir de procesos tectónicos durante el Cretácico (Conabio, 2010). En el área de estudio domina el tipo de roca ígnea extrusiva, donde la unidad geomorfológica comprende laderas de montaña con pendientes de mediana a muy fuertemente inclinadas (Rodríguez-González, 2015). Los tipos de suelo dominantes son Cambisol y Regosol (INEGI, 2015). Presenta un clima templado subhúmedo, con una temperatura media anual de 14.2 °C. La temperatura media del mes más cálido (mayo y junio) es de 16.7 °C, mientras que la temperatura media del mes más frío es 10.8 (enero y febrero), con una precipitación promedio anual de 2,003 mm con lluvias en verano. La precipitación está concentrada entre mayo a septiembre (Estación 14271, Servicio Meteorológico Nacional, Conagua, 2025). Las principales actividades productivas son la agricultura de temporal, con apertura de áreas para el establecimiento de pastizales, la minería y la extracción de madera (Vargas-Rodríguez et al., 2010).
El bosque de Pinus–Quercus–Abies es una asociación vegetal con distribución restringida cerca del límite altitudinal de la vegetación arbórea (Cuevas-Guzmán et al., 2011; del Castillo-Batista et al., 2018). En diversos estudios florísticos de la zona se ha encontrado una alta diversidad florística para numerosas formaciones vegetales de alta montaña, entre ellas el bosque de Pinus con diferentes asociaciones vegetales (Cuevas-Guzmán et al., 2011; Guerrero-Hernández et al., 2014, 2019). Los tipos de bosque de alta montaña son representados principalmente por bosques de coníferas, donde los géneros representativos son Pinus, Abies, Cupressus, Juniperus y algunas latifoliadas como Quercus, Alnus y Arbutus (Perry et al., 1998).
La diversidad florística del área de estudio corresponde a un tipo de vegetación dominado en el estrato arbóreo por Abies jaliscana, A. religiosa, Alnus acuminata, A. jorullensis, Arbutus xalapensis, Carpinus tropicalis, Clethra fragrans, Pinus devoniana, P. douglasiana, P. oocarpa, P. jaliscana, P. herrerae, P. pseudostrobus, Podachaenium eminens, Quercus castanea, Q. calophylla, Q. scytophylla, Q. magnoliifolia, Q. obtusata y Ternstroemia lineata. Se encuentran algunos arbustos dominantes como Archibaccharis serratifolia, Baccharis heterophylla, B. pteronioides, Fuchsia microphylla, Mimosa albida, Monnina ciliolata y Salvia iodantha. El componente herbáceo es diverso, como Amaranthus palmeri, Chenopodium ambrosoides, Iresine diffusa, Antigonon flavescens, Eryngium alternatum, Bidens odorata, B. aurea, Dahlia coccinea, Ageratina choricephala, Fuchsia fulgens, Rumfordia floribunda, Melampodium perfoliatum, Piqueria triflora, Tagetes filifolia, T. lucida, Crotalaria filifolia, C. mollicula, Dalea obreniformis. Por lo que en este estudio nos referimos a este tipo de vegetación como bosque de Pinus–Quercus–Abies (BPQA) por ser los taxones dominantes.
Se realizó un inventario de la vegetación leñosa en 14 parcelas circulares de 500 m2 cada una, siguiendo el protocolo estándar de Olvera-Vargas y Figueroa-Rangel (2023). Cada parcela se ubicó a una distancia mínima de 50 m respecto de las demás. En cada una se registraron datos ambientales y de localización (localidad, coordenadas, elevación, exposición y pendiente) (material suplementario: tabla S1). Para todos los taxones leñosos con diámetros ≥ 2.5 cm arriba de 1.30 m del nivel del suelo (DN). Con estas mediciones se estimaron indicadores estructurales a nivel de parcela, incluyendo densidad (individuos ha-1) y área basal (m2 ha-1). Se recolectaron ejemplares de herbario para confirmar la determinación taxonómica de las especies. La clasificación de familias siguió el sistema Angiosperm Phylogeny Group IV (2016) y la actualización nomenclatural se verificó en la base de datos Tropicos (www.tropicos.org). Este diseño y conjunto de métricas permiten caracterizar de manera comparable la estructura y composición de la vegetación en el área de estudio.
Se recolectaron 22 muestras de polen moderno dentro de las parcelas para evaluar la relación polen-vegetación a escala local (tabla S1). En cada parcela se establecieron subparcelas de 1 m2, dentro de las cuales se obtuvieron muestras de sedimento superficial (suelo y musgo) a una profundidad de 0-5 cm (Hjelle, 1999). Cada muestra se almacenó para su posterior extracción y análisis. Dado que los ambientes de depósito reciben polen de fuentes locales y regionales, el espectro palinológico está condicionado por procesos tafonómicos modulados por el clima y las propiedades del suelo, lo que afecta la relación polen-vegetación observada (Jacobson y Bradshaw, 1981). En este contexto, asumimos que la señal obtenida es representativa del ecosistema fuente a la escala de captura de nuestros sustratos. Si bien superficies extensas como lagos y humedales maximizan el componente regional y facilitan el vínculo con paleorregistros (Markgraf et al., 2002; Ortuño et al., 2011), el muestreo de sedimento superficial aporta la resolución local necesaria para calibrar e interpretar esa señal más integrada en el sitio de estudio.
Con la finalidad de comparar la relación entre el polen moderno y fósil, se extrajo un núcleo de sedimento que abarca los últimos 1,580 años en el BPQA (del Castillo-Batista et al., 2018). El muestreo consistió en la extracción de un núcleo de sedimento de 62 cm de profundidad con un taladro Eijelkamp (“forest hollow”, sensu Calcote 1995, 1998), en un área abierta del bosque a 2,119 m de elevación en la localidad de La Cumbre de Guadalupe (20°10’17.82” N, 104°42’42.32” O) (fig. 1). Se tomaron muestras de sedimento cada 1 cm para su análisis. La cronología se determinó con base en 4 fechados radiocarbono 14C por medio de la técnica AMS (por sus siglas en inglés, accelerator mass spectrometer) en el laboratorio Beta Analytic, Miami, Florida. Para la calibración en años calendáricos fueron utilizados los programas INCAL04.14 y CALIB v.5.02 de Stuiver y Reimer (Stuiver y Reimer, 1993). Para elaborar la cronología se obtuvo un modelo edad-profundidad por interpolación lineal de acuerdo con el método de Bennett (1994), con el apoyo del software Psimpoll (v. 4.25), lo que permitió estimar la tasa de sedimentación promedio de 25 años/cm.
Para la extracción de polen moderno y el polen fósil de las secuencias sedimentarias, se siguió un protocolo estándar de acetólisis Bennett y Willis (2001). Se agregó una pastilla de Lycopodium clavatum L. (Batch 3862) a cada muestra con el fin de estimar la concentración de granos de polen por cm3 (Stockmarr, 1971). Para obtener el tamaño de muestra estadísticamente significativo, se contabilizaron 400 granos de polen por muestra (Maher, 1972). La determinación de polen y esporas se realizó utilizando la colección de referencia palinológica “The Mexican reference collection of Global Pollen Project” (Figueroa-Rangel, 2017), además de claves palinológicas y bibliografía especializada. Para armonizar la resolución taxonómica entre los ensambles de polen fósil y moderno (en los que no siempre es posible asignar el nivel taxonómico de especie) se elaboró una tabla de equivalencias (material suplementario: tabla S2) que vincula cada tipo de polen moderno con sus especies vegetales potencialmente representadas, agrupadas por forma biológica. Los conteos de polen se expresaron como la suma total en porcentaje del polen de árboles, arbustos, herbáceas y esporas de Pteridophyta sensu lato, consideradas en este trabajo como abundancia de polen (Bennett y Willis, 2001). El polen desconocido se contó, pero fue excluido de la suma total de polen en ambos ensambles. El diagrama de polen fósil se elaboró en Psimpoll (v. 4.25) (Bennett, 2005), mientras que, el de polen moderno en el programa Tilia Graph (Grimm, 2020).
La contribución de cada una de las especies encontradas en la vegetación se estimó utilizando el índice de valor de importancia (IVI) para árboles (Matteucci y Colma, 1982; Whittaker, 1967). Este índice es un valor ponderado de la estructura del bosque, se calcula sumando las variables estructurales de la densidad relativa (número de árboles del taxón dividido entre el número total de árboles de todos los taxones), la frecuencia relativa (número de muestras en las que el taxón aparece dividido entre el número total de muestras) y el área basal relativa (área basal total del taxón dividida entre el área basal total de todos los taxones) (Cottam y Curtis, 1956). De esta manera, el IVI proporciona una medida de la importancia relativa de cada especie en el rodal. Así mismo, para el componente arbóreo se estimó la media de diámetro y altura por especie de todas las parcelas muestreadas.
Para describir la distribución de los diámetros y las alturas de las especies arbóreas en las parcelas muestreadas, se estimó la media del diámetro a la altura de 1.30 m (DN) y la altura total por especie. Se utilizaron diagramas de caja y bigote para identificar la mediana, los cuartiles (Q1 y Q3), el rango intercuartílico y la presencia de valores atípicos, proporcionando información sobre la variabilidad y la tendencia central de cada variable. El análisis fue realizado mediante el software estadístico R Project 3.6.1 (Ihaka y Gentleman,1993).
Para analizar la relación entre las especies y las parcelas basados en el método de Olvera-Vargas y Figueroa-Rangel (2023) se realizó un Análisis de Conglomerados. Es una técnica estadística multivariada que tiene como objetivo agrupar elementos o variables para lograr la máxima homogeneidad dentro de cada grupo y la mayor diferencia entre los grupos. Se trata de una clasificación jerárquica aglomerativa (Clarke y Gorley, 2015). Se utilizó la matriz de densidad de las especies, la cual fue transformada con una doble raíz cuadrada para mejorar la ponderación de las especies raras y abundantes. El análisis se realizó utilizando la medida de distancia Bray-Curtis y como método de unión de grupos de los promedios ponderados, con el apoyo del software PRIMER V.7 (Clarke y Gorley, 2015).
Utilizando datos de presencia-ausencia del polen moderno y de la vegetación, se calcularon 3 índices de asociación R (Davis, 1984), el índice de asociación (A), el índice de sobrerrepresentación (O) y el índice de subrepresentación (U). Estos índices permiten determinar si los taxones están presentes de manera simultánea, tanto en el conjunto de polen como en la vegetación (León-Carreño et al., 2019). Se calcularon con las fórmulas que se describen a continuación:
A = B0 / P0 + P1 + B0
O = P0 / P0 + B0
U = P1 / P1 + B0
Donde B0 corresponde al número de parcelas en las que tanto el tipo polínico como el grupo vegetal asociado están presentes, P0 representa el número de parcelas en las que se registra el polen, pero la planta que lo produce no está presente en la vegetación, y P1 es el número de parcelas en las que el tipo de polen está ausente pero el taxón vegetal se encuentra presente en la vegetación. Los valores del índice de asociación varían entre 0 y 1; donde A = 1 indica que el tipo polínico y el taxón vegetal están siempre presentes y si A = 0, alguno de los 2 elementos se encuentra ausente en todos los levantamientos. Posteriormente, los valores de los índices de cada taxón se agruparon en las siguientes categorías: 1) tipo fuertemente asociado “TFA” cuando A > 0.65; 2) tipo asociado “TA”, cuando A varía entre 0.65 y 0.5; 3) tipo débilmente asociado “TDA”, donde A < 0.5 y además O y U son positivos; 4) tipo sobrerrepresentado “TOR”, cuando A < 0.5 y U = 0; 5) tipo subrepresentado “TUR”, con A < 0.5 y O = 0, y 6) tipo no asociado “TNA”, en donde A = 0, además O y U son positivos (Fjordheim et al., 2018).
La diversidad de taxones de los datos de polen moderno y fósil se estimó mediante la diversidad de Hill N0 sumando el número de taxones identificados en cada conjunto (Hill, 1973). Se calcularon los valores de diversidad de Hill N1 para obtener más información sobre las características de diversidad del polen. Esta métrica considera la riqueza de taxones en cada muestra y pondera cada taxón según su abundancia relativa. Asimismo, se seleccionó el número de Hill N2 para representar el número de taxones de polen fósil que son muy abundantes (dominantes) (Figueroa-Rangel et al., 2020). Estos cálculos se realizaron utilizando la librería “vegan” (Oksanen et al., 2010), disponible en el paquete estadístico R-v3.4.3 (R Core Team, 2019).
Con el propósito de comparar la composición taxonómica entre los ensambles de polen fósil con los de polen moderno, se aplicó la función de comparación de análogos múltiples (analog.mult), que calcula índices de distancia o disimilitud entre las muestras de polen fósil con el moderno. Para ello se utilizó la distancia cordal (square chord distance en inglés) mediante la librería “paleoMAS” (Correa-Metrio et al., 2015) disponible en R (R Core Team, 2019).
Resultados
Se registraron 6 familias, 8 géneros y 17 especies en las 14 parcelas analizadas. Las familias con más riqueza de especies fueron Pinaceae y Fagaceae. Los géneros más frecuentes presentes en las parcelas fueron: Pinus–Quercus–Abies. Pinus pseudostrobus fue la especie dominante en el dosel del rodal por su IVI (tabla 1), seguida por Abies jaliscana, Pinus devoniana, Quercus castanea y P. herrerae, todos ellos identificados como elementos importantes de acuerdo con su IVI. Por otro lado, Q. scytophylla, Q. magnoliifolia y P. douglasiana presentan los valores más bajos de IVI.
El análisis de cuartiles de los diámetros normales de las especies arbóreas muestra que la mayoría de los individuos presentan diámetros inferiores a 60 cm (fig. 2a). Se observa una notable dispersión en la mediana de Vachellia farnesiana, lo que indica una alta variabilidad en el diámetro de esta especie. En contraste, Baccharis pteronioides, Monnina ciliolata y Rumfordia floribunda presentan diámetros más bajos y menos dispersos. Además, Pinus devoniana, P. douglasiana y P. pseudostrobus concentran los valores de diámetros más altos entre las especies analizadas. En cuanto a las alturas, la mayoría de las especies registran valores inferiores a los 30 m (fig. 2b). Quercus calophylla se destaca por su variabilidad en la distribución de alturas, mientras que Baccharis pteronioides, Monnina ciliolata y Rumfordia floribunda presentan alturas significativamente menores y con menor dispersión.
La clasificación de las parcelas generó 4 grupos, mismos que se formaron para la clasificación de las especies. Cuando se presenta el resultado a través del dendrograma de sombras, se observó que algunas especies como Pinus pseudostrobus y P. devoniana se encuentran en los 4 grupos de parcelas, pero en general, hay coincidencia de grupos de especies con los grupos de parcelas. Abies religiosa, Rumfordia floribunda, Monnina ciliolata y Arbutus xalapensis presentaron sus mayores abundancias en la parcela P10 y P11, mientras que Quercus herrerae, Q. magnolifolia y Q. castanea tienen su mayor expresión en abundancias en las parcelas P01, P02, P07 y P08 (fig. 3).
En el espectro polínico del bosque se identificaron 41 taxones, pertenecientes a 31 familias y 40 géneros, que incluyen árboles, arbustos, herbáceas y helechos (material suplementario: tabla S3). Pinus se registró como el taxón dominante en el estrato arbóreo, con más de 60% del espectro polínico. De manera similar, los taxones de Quercus (< 20%) y Abies (< 15%) se encontraron en todo el espectro polínico con porcentajes considerables. En el estrato herbáceo, Poaceae se registró como la más abundante, se encontró en todas las muestras con un porcentaje superior a 5%. Los taxones de Eupatorium (> 5%) y Baccharis (> 5%) se registraron con porcentajes bajos en el componente herbáceo. Polypodiaceae prevaleció con un porcentaje mayor a 5%. No obstante, el grupo de los helechos no alcanzó una representación importante en el espectro polínico (fig. 4).
El índice de Davis revela que la mayoría de los taxones polínicos se adscriben al grupo de tipo asociado (TOR); los taxones de polen de Pinus, Quercus, Abies, Arbutus xalapensis, Baccharis pteronioides y Rumfordia floribunda, pertenecen al grupo de tipo asociado (TA), mientras que Monnina ciliolata es el único que se encuentra en el tipo subrepresentado (TUR). Ningún taxón demostró ser fuertemente asociado (TFA) (tabla 2).
Tabla 1
Valores relativos para densidad (Dr), área basal (ABr) y frecuencia (Fr) en árboles del BPQA. Índice de valor de importancia (IVI) para árboles y arbustos.
| Familias | Especies | Dr (%) | ABr (%) | Fr (%) | IVI (%) |
| Asteraceae | Baccharis pteronioides (Bacpte) | 1.02 | 0.02 | 4.23 | 1.86 |
| Rumfordia floribunda (Rumflo) | 15.90 | 0.39 | 4.23 | 6.84 | |
| Ericaceae | Arbutus xalapensis (Axal) | 0.76 | 4.12 | 4.05 | 8.94 |
| Fabaceae | Vachellia farnesiana (Vacfar) | 0.76 | 2.80 | 1.41 | 1.66 |
| Fagaceae | Quercus calophylla (Quecal) | 1.53 | 3.48 | 5.63 | 3.54 |
| Quercus castanea (Quecas) | 5.98 | 12.84 | 9.72 | 9.51 | |
| Quercus magnoliifolia (Quemag) | 1.91 | 1.10 | 4.23 | 2.41 | |
| Quercus obtusata (Queobt) | 5.60 | 6.65 | 8.45 | 6.9 | |
| Quercus scytophylla (Quescy) | 0.89 | 0.34 | 1.41 | 0.87 | |
| Pinaceae | Abies jaliscana (Abijal) | 14.89 | 12.97 | 12.16 | 13.33 |
| Abies religiosa (Abirel) | 7.38 | 2.13 | 4.23 | 4.58 | |
| Pinus devoniana (Pindev) | 9.29 | 14.61 | 11.27 | 11.72 | |
| Pinus douglasiana (Pindou) | 1.65 | 3.08 | 2.82 | 2.51 | |
| Pinus herrerae (Pinher) | 5.47 | 11.45 | 8.45 | 8.46 | |
| Pinus oocarpa (Pinooc) | 4.20 | 2.80 | 5.63 | 4.21 | |
| Pinus pseudostrobus (Pinpse) | 22.14 | 21.19 | 11.27 | 19.19 | |
| Polygalaceae | Monnina ciliolata (Moncil) | 0.64 | 0.02 | 1.41 | 0.69 |
El diagrama polínico del bosque fue dividido en 3 zonas palinológicas mediante un análisis de conglomerados (CONISS) (fig. 5). A lo largo de toda la secuencia del ensamble de polen fósil, se observó un patrón heterogéneo caracterizado por la oscilación de los taxones leñosos herbáceos y helechos. La primera zona palinológica (C-1) corresponde al periodo de 1,580-1,550 años cal AP y muestra una mayor abundancia de taxones leñosos como Alnus, Pinus y Quercus. En cuanto a los taxones herbáceos, los más abundantes fueron Asteraceae, Cyperaceae y Poaceae. La segunda zona (C-2) corresponde al periodo de 1,550-740 años cal AP, abarca el periodo más extenso, presenta un patrón similar al anterior, con la adición de taxones de Fabaceae y Rosaceae. La tercera zona (C-3a) corresponde al año 740-700 años cal AP, se observa un patrón divergente caracterizado por una disminución de Aspleniaceae y Bromeliaceae. En los últimos 700-100 años cal AP (zona C-3b) se aprecia un patrón heterogéneo con oscilaciones en los porcentajes de Pinus, Quercus y Alnus en el grupo de los taxones leñosos, así como en los de Asteraceae, Cyperaceae y Poaceae en el grupo de las herbáceas. En contraste, las esporas como Hymenophyllaceae, Polypodiaceae y Sphagnum disminuyen considerablemente sus porcentajes (fig. 5).
La diversidad de Hill para los conjuntos de polen moderno para el orden N0, indicó que la riqueza es de aproximadamente 16 taxones (con un intervalo de confianza de 95%). Para el orden N1 y N2 obtuvo valores de 3.79-4.76 y 2.4-2.9, respectivamente. Para el conjunto de polen fósil, N0 indicó que la riqueza es de alrededor de 34 taxones (con un intervalo de confianza de 95%), mientras que para N1 y N2 fueron de 14.4-15.8 y 9.4-10.5, respectivamente (figs. 4, 5).
La comparación entre los ensambles de polen moderno y la secuencia de polen fósil de La Cumbre de Guadalupe (fig. 6) mostró distancias cordales al cuadrado que variaron aproximadamente de 0.30 a 0.72. A lo largo de la cronosecuencia predominaron valores intermedios (≈ 0.45-0.60). Se identificaron mínimos locales de disimilitud (˂ 0.40) en intervalos temporales estrechos, destacando un pulso centrado alrededor de ~ 850 años cal AP. En contraste, se observaron franjas con mayor disimilitud (≥ 0.65) especialmente después de ~ 1,000 años cal AP y hacia los extremos de la secuencia. El mejor análogo moderno por nivel fósil rara vez presentó distancias ≤ 0.35 y no se registraron valores cercanos a 0; por lo tanto, no se detectaron análogos modernos muy cercanos para los niveles analizados.
Discusión
El análisis de la vegetación actual en la sierra de Cacoma, Jalisco, muestra que la composición florística está dominada por especies de Pinus–Quercus-Abies como los elementos estructurales más importantes según el IVI. Estos resultados coinciden con estudios previos que han identificado a estos taxones como dominantes en los bosques templados del occidente de México (Cuevas-Guzmán et al., 2011; Guerrero-Hernández et al., 2014, 2019). La estructura diamétrica de la comunidad sugiere una comunidad vegetal joven, con la mayoría de los individuos de diámetros menores de 60 cm, lo que coincide con patrones observados en otras regiones montañosas donde la explotación selectiva ha modificado la distribución diamétrica de las especies arbóreas (Guerrero-Hernández et al., 2022).
Tabla 2
Índice de Davis. La columna de taxones muestra el nombre científico de las especies. (A) significa la asociación, (O) sobrerrepresentación y (U) subrepresentación de los taxones registrados en los ensambles polínicos, basados en datos de presencia/ausencia, tanto de lluvia de polen como de la vegetación del BPQA. Los índices de asociación se clasificaron en: TFA = tipos fuertemente asociados, TA = tipos asociados, TDA = tipos débilmente asociados, TOR = tipos sobrerrepresentados, TUR = tipos subrepresentados y TNA = tipos no asociados.
| Taxones | A | O | U | Tipo |
| Pinus | 0.50 | 0.33 | 0.33 | TA |
| Quercus | 0.50 | 0.33 | 0.33 | TA |
| Abies | 0.50 | 0.35 | 0.32 | TA |
| Alnus acuminata | 0.50 | 0.50 | 0.00 | TOR |
| Alnus jorullensis | 0.50 | 0.50 | 0.00 | TOR |
| Arbutus xalapensis | 0.50 | 0.40 | 0.25 | TA |
| Cestrum | 0.50 | 0.50 | 0.00 | TOR |
| Cornus disciflora | 0.50 | 0.50 | 0.00 | TOR |
| Eugenia | 0.50 | 0.50 | 0.00 | TOR |
| Ficus | 0.50 | 0.50 | 0.00 | TOR |
| Fraxinus uhdei | 0.50 | 0.50 | 0.00 | TOR |
| Juglans major | 0.50 | 0.50 | 0.00 | TOR |
| Magnolia pacifica | 0.50 | 0.50 | 0.00 | TOR |
| Ostrya-carpinus | 0.50 | 0.50 | 0.00 | TOR |
| Ternstroemia lineata | 0.50 | 0.50 | 0.00 | TOR |
| Tilia mexicana | 0.50 | 0.50 | 0.00 | TOR |
| Zinowiewia concinna | 0.00 | 0.00 | 0.00 | – |
| Cupressaceae | 0.00 | 0.00 | 0.00 | – |
| Vachellia farnesiana | 0.50 | 0.50 | 0.00 | TOR |
| Baccharis pteronioides | 0.50 | 0.38 | 0.27 | TA |
| Rumfordia floribunda | 0.50 | 0.25 | 0.40 | TA |
| Monnina ciliolata | 0.50 | 0.00 | 0.50 | TUR |
El análisis del dendrograma de sombras identificó 4 grupos de especies y 4 de parcelas, evidenciando patrones de distribución diferenciados dentro del paisaje forestal. La estructura de los grupos sugiere coexistencia mediada por diferencias fisiológicas y ambientales. El género Pinus domina en la mayoría de los conjuntos en congruencia con su especialización en hábitats de fertilidad baja a moderada y en condiciones abiertas impuestas por aridez, frío o disturbio, así como con historias evolutivas asociadas con altas frecuencias de perturbación, en particular el fuego (Keeley y Zedler, 1998). En contraste, Quercus y Abies conforman subgrupos bien definidos que reflejan preferencias por microambientes más fríos y húmedos (Cuevas-Guzmán et al., 2011; Guerrero-Hernández et al., 2022). La alta similitud florística entre algunas parcelas sugiere que existen factores ambientales y de manejo forestal que han promovido la homogeneización de la vegetación en ciertas áreas, fenómeno que ha sido documentado en otros estudios de dinámica de bosques templados mexicanos (Guerrero-Hernández et al., 2019).
La composición del ensamble polínico moderno refleja la estructura y composición del bosque montano dominado por Pinus–Quercus–Abies. La alta representación de Pinus en el espectro polínico (> 60%) es consistente con estudios previos que han demostrado su alta producción polínica y dispersión anemófila en ecosistemas templados (Castro-López et al., 2020), y coinciden con estudios de lluvia polínica en bosques del occidente y centro de México (Figueroa-Rangel et al., 2016; Lozano-García et al., 2014), los cuales reportan estos elementos con porcentajes superiores a 40% en Pinus y Quercus, mientras que Abies oscila entre 10 y 20%. Estos resultados se explican por los mecanismos de dispersión anemófila y por la gran cantidad de granos de polen que producen (Faegri e Iversen, 1989). La presencia de Quercus y Abies con valores inferiores a Pinus, pero aún representativos (< 20% y < 15%, respectivamente), coincide con su menor producción polínica y las características de dispersión de estos géneros, factores que afectan su representatividad en los registros polínicos modernos (Rodríguez-Pérez et al., 2025).
En cuanto a la vegetación herbácea, el predominio de Poaceae (> 5%) indica su importancia en los espacios abiertos del bosque y su alta eficiencia en la producción y dispersión del polen (Hjelle, 1998). La baja representación de Eupatorium y Baccharis (< 5%) sugiere que estas especies, aunque presentes en la vegetación actual, tienen una menor contribución al espectro polínico debido a sus estrategias reproductivas y mecanismos de dispersión, como la entomofilia (Faegri e Iversen, 1989). De manera similar, las esporas de Polypodiaceae, aunque registradas con valores > 5%, no alcanzaron una representación dominante en el espectro polínico, lo que puede atribuirse a que a medida que aumenta la altitud, la representación del polen de la vegetación herbácea disminuye (Markgraf, 1980).
El análisis de representación polínica a través del índice de Davis confirma que los taxones dominantes en el bosque se encuentran principalmente en el grupo de tipo asociado (TA), lo que sugiere una correspondencia adecuada entre la vegetación y el espectro polínico. El predominio del grupo sobrerrepresentado (TOR) para la mayoría de los taxones indica que ciertos elementos del bosque reproducen y dispersan polen en proporciones superiores a su biomasa real en el ecosistema (Jackson y Overpek, 2000). Monnina ciliolata, identificado como tipo subrepresentado (TUR), refleja una producción polínica baja por sus mecanismos de dispersión especializados, lo que puede dificultar su identificación en registros fósiles (Faegri, 1966). La ausencia de taxones en el grupo fuertemente asociado (TFA) sugiere que ningún elemento arbóreo o herbáceo presenta una correspondencia entre su presencia en el bosque y su representación polínica. Este patrón puede explicarse por síndromes de polinización especializados, bajas abundancias locales (León-Carreño et al., 2019) una estructura poblacional dominada por individuos juveniles con pocos adultos reproductivos, como sugiere la distribución de diámetros y alturas (fig. 2).
El ensamble polínico fósil refleja la dinámica de la vegetación durante los últimos ~ 1,580 años, que evidencia cambios en la composición florística asociados a variaciones climáticas y potenciales impactos antrópicos. El diagrama polínico muestra 3 zonas diferenciadas, que sugiere una transición en la estructura del bosque, desde una composición dominada por taxones leñosos hasta un escenario más heterogéneo en los últimos siglos. En la primera zona palinológica (C-1), la dominancia de Alnus, Pinus y Quercus sugiere la prevalencia de un bosque templado de montaña con condiciones húmedas, lo que concuerda con estudios previos en la región del occidente de México (Figueroa-Rangel et al., 2016, 2020). La alta representación de Asteraceae, Cyperaceae y Poaceae indica la presencia de claros en el dosel forestal y espacios abiertos con vegetación herbácea.
Durante la segunda fase, que abarca el periodo más extenso (zona C-2), se mantiene la estructura del bosque, pero con la incorporación de taxones como Fabaceae y Rosaceae, lo que podría indicar una diversificación de la vegetación secundaria y la posibilidad de fluctuaciones climáticas que favorecieron la heterogeneidad del ecosistema (Arellano, 2024). La estabilidad relativa de los elementos leñosos indica que, bajo condiciones climáticas favorables que sostienen la cobertura forestal, el bosque mantiene su estructura y composición con cambios mínimos, evidenciando una alta capacidad de persistencia frente a posibles perturbaciones de baja a moderada intensidad.
El cambio más significativo ocurre en la tercera zona (C-3a), donde se observa una disminución de Aspleniaceae y Bromeliaceae, lo que podría estar vinculado a un enfriamiento o a una reducción en la humedad ambiental, en concordancia con el evento de la Pequeña Edad de Hielo documentado en bosques de alta montaña del occidente de México (Figueroa-Rangel et al., 2019, 2020). Posteriormente, en los últimos 700 años (zona C-3b), la heterogeneidad en las oscilaciones de Pinus, Quercus y Abies sugiere periodos de regeneración y disturbio en el bosque, posiblemente por influencia de actividad humana (del Castillo et al., 2018). La disminución de Cyperaceae, Hymenophyllaceae, Polypodiaceae y Sphagnum podría estar relacionada con una menor disponibilidad de ambientes húmedos o la alteración de los suelos debido a disturbios naturales o inducidos.
El análisis polínico moderno confirma la composición arbórea identificada en campo, con la dominancia de Pinus, Quercus y Abies. Sin embargo, la diversidad de Hill indica una riqueza relativamente baja en el polen moderno (N0 = 16, fig. 4) en comparación con el registro fósil (N0 = 34, fig. 5), lo que sugiere una composición más diversa de la vegetación en el pasado. Esta diferencia podría atribuirse a la dispersión diferencial del polen, a sesgos tafonómicos en la preservación en el sedimento y a la representatividad desigual de ciertos taxones en el espectro polínico (Birks et al., 2011; Xiao et al., 2016). Estudios similares han reportado una disminución de la diversidad arbórea en el Holoceno tardío en otras zonas templadas y tropicales de México, vinculada a eventos de cambio climático y actividades antrópicas (Lozano-García et al., 2021; Ortega-Rosas et al., 2008).
El análisis de los ensambles polínicos modernos y fósiles revela patrones temporales de similitud y disimilitud que sugieren cambios en la composición y estructura de la vegetación a lo largo de los últimos ~ 1,580 años. La mayor similitud entre los registros modernos y fósiles se presenta entre ~ 850 y 400 años cal AP, lo que es congruente con lo reportado en la sierra de Manantlán alrededor de ~ 850 años cal AP, caracterizado por la disminución de Pinus y el aumento de Quercus y Alnus, lo que pareciera una contracción del bosque dominado por Pinus y la expansión relativa de latifoliadas (Figueroa-Rangel et al., 2016). Este intervalo coincide con la Anomalía Climática Medieval (ACM) (800-1200 d.C.), caracterizada regionalmente por condiciones más cálidas y en diversos registros, mayor aridez estacional congruentes con otras secuencias polínicas del occidente y centro de México (Lachniet et al., 2012; Metcalfe et al., 2010), donde Quercus y Alnus se han propuesto como taxones indicadores de sequía (Lozano-García et al., 2021).
La similitud observada hacia ~ 400 años cal AP se acompaña de una expansión relativa de Poaceae y Asteraceae, familias típicamente pioneras y asociadas a ambientes abiertos o recientemente perturbados. Este intervalo coincide con la Pequeña Edad de Hielo (LIA) (1300 y 1850 d.C.), un periodo más frío que el actual que pudo reducir la productividad del bosque, favoreciendo claros y paisajes abiertos propicios para el establecimiento de pastizales (Bradley y Jones, 1993; Mann et al., 2009). Adicionalmente, las perturbaciones antrópicas de los últimos siglos (cambios de uso de suelo, tala y fuego) posiblemente reforzaron esa apertura del paisaje, lo que incrementó la representación polínica de Poaceae y Asteraceae (Lozano-García et al., 2021).
Por otro lado, los periodos de mayor disimilitud identificados hace ~ 1,500, 1,000 y 300 años cal AP, sugieren episodios de cambio ambiental y posibles alteraciones en la vegetación local. Estos periodos de disimilitud se relacionan estrechamente con las 3 zonas identificadas previamente mediante CONISS. La disimilitud hace ~ 1,500 años cal AP coincide con la transición entre las zonas más antiguas de la secuencia, caracterizada por una menor representación de Pinus y un mayor aporte de taxones leñosos y herbáceos indicativos de condiciones más húmedas, lo que significa mayor heterogeneidad ambiental y cambios en la dinámica del bosque (Lachniet et al., 2012). Asimismo, la disimilitud hace ~ 1,000 años cal AP se asocia con un incremento de Abies y una reducción de taxones de ambientes abiertos, lo cual lleva a una reorganización en la composición del bosque, posiblemente asociada a variaciones climáticas durante la Anomalía Climática Medieval; estos episodios han sido relacionados con fluctuaciones de humedad y temperatura que afectaron la distribución de especies arbóreas (Figueroa-Rangel et al., 2016; Metcalfe et al., 2010). En el intervalo más reciente, aproximadamente los últimos 300 años cal AP, se observa un aumento de la disimilitud que marca la transición hacia una señal polínica moderna claramente diferenciada de los ensambles fósiles. Esta modificación puede estar asociada, por un lado, con la variabilidad climática vinculada a la Pequeña Edad de Hielo y por otro lado, con cambios de origen antropogénico posteriores a la colonización española, que implicaron modificaciones del uso del suelo y una mayor presión sobre los ecosistemas como los incendios y procesos de deforestación, que alteraron la estructura y composición del bosque y generaron ensambles polínicos distintos respecto de las condiciones previas (Lozano-García et al., 2021). Estos hallazgos resaltan la interconexión entre los procesos ecológicos y climáticos en la configuración de los ecosistemas forestales, y proporcionan evidencia sobre la sensibilidad de estos ecosistemas a los cambios ambientales pasados.
El análisis de los ensambles polínicos modernos y fósiles en la sierra de Cacoma, Jalisco en particular en la Cumbre de Guadalupe, ha permitido reconstruir la dinámica de la vegetación durante los últimos ~ 1,580 años, que evidencian cambios en la composición y estructura del bosque en respuesta a factores climáticos y antrópicos. Los resultados mostraron que la vegetación actual está dominada por Pinus pseudostrobus y Abies jaliscana, los elementos estructurales más relevantes. Esta composición coincide con la representación polínica moderna, donde Pinus–Quercus–Abies fueron los taxones dominantes, lo que refleja la relación entre el ensamble polínico y la estructura del bosque. Sin embargo, el registro polínico fósil indica que la vegetación del pasado presentaba una mayor diversidad de especies, lo que sugiere una mayor heterogeneidad de la vegetación en el pasado. La disminución de la diversidad efectiva en los últimos siglos podría estar relacionada con el impacto de eventos climáticos, como la Pequeña Edad de Hielo (~ 850-400 AP), así como la influencia de actividades antrópicas, como el cambio en el uso de suelo, los incendios y la explotación forestal.
El análisis de disimilitud entre ensambles de polen moderno y fósil identificó intervalos de mayor similitud estructural del bosque (~ 850 y 400 cal AP), intercalados con episodios de cambio (~ 1,500, ~ 1,000 y ~ 300 cal AP). Estos periodos coinciden con transiciones climáticas del Holoceno tardío, incluida la Anomalía Climática Medieval y la Pequeña Edad de Hielo, que modularon la composición de las comunidades vegetales. Asimismo, la evidencia indica que en los últimos siglos, la actividad humana (incendios, deforestación) ha alterado la estructura forestal y ha favorecido la homogeneización de la vegetación en algunas zonas.
Los hallazgos señalan la relevancia del análisis palinológico en la reconstrucción de la historia ambiental y ecológica de los bosques montanos del occidente de México. Futuros estudios podrían integrar un análisis multiproxy, como datos isotópicos y modelado climático, para mejorar la resolución temporal de los cambios en la vegetación y su relación con la variabilidad climática. Asimismo, se podrían desarrollar análisis de correspondencia entre los conjuntos de polen moderno y la vegetación actual para identificar los factores que explican la ausencia de ciertos taxones esperados, a pesar de la evidencia que indica una correspondencia entre la composición de polen local y la vegetación del bosque.
Agradecimientos
Agradecemos el apoyo financiero de los proyectos Conahcyt CB-2008-106435 y UDG-PTC-1403. Asimismo,
agradecemos al Laboratorio de Paleoecología y Cambio Climático del Departamento de Ecología y Recursos Naturales del Centro Universitario de la Costa Sur, Universidad de Guadalajara, y al Laboratorio de Paleoecología, Paleoclimatología y Cambio Climático del Departamento de Dinámica Superficial del Instituto de Geología, UNAM, por las facilidades, el apoyo logístico y la infraestructura brindados para el desarrollo del estudio. El acceso a la zona de estudio fue brindado por los habitantes de la comunidad de la Cumbre de Guadalupe. También expresamos nuestro reconocimiento a quienes colaboraron en el trabajo de campo, así como a Carlos Armando Pacheco Contreras por la elaboración del mapa y el trabajo de campo. Finalmente, agradecemos a los revisores anónimos por sus observaciones y sugerencias que contribuyeron a mejorar este trabajo.
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Survival, growth and sexual dimorphism of the land crab Cardisoma guanhumi under semi-controlled captive conditions
Supervivencia, crecimiento y dimorfismo sexual del cangrejo terrestre Cardisoma guanhumi en condiciones semicontroladas de cautiverio
Karla Verónica Santana-Aguayo, Carlos Iván Pérez-Rostro *, Martha Patricia Hernández-Vergara
Instituto Tecnológico de Boca del Río, División de Estudios de Posgrado e Investigación, Laboratorio de Mejoramiento Genético y Producción Acuícola, Km. 12 Carretera Veracruz–Córdoba, 94290 Boca del Río, Veracruz, Mexico
*Corresponding author: ivandna02@hotmail.com (C.I. Pérez-Rostro)
Received: 21 April 2025; accepted: 23 February 2026
Abstract
This study evaluated the survival, growth, and reproductive performance of Cardisoma guanhumi maintained for 12 months under semi-controlled conditions. A total of 341 adult crabs were collected from mangrove habitats and housed in a naturalized enclosure equipped with buried PVC shelters, freshwater and saltwater basins, and automated sprinklers to maintain humidity. Crabs were fed a plant-based diet supplemented with tilapia muscle and monitored monthly. Final survival reached 76%, and 68.6% of females became ovigerous, supporting the species’ reproductive viability under captive conditions. Morphometric growth was recorded in 170 intact individuals. Notably, carapace width increased by approximately 1 mm/month despite the near absence of visible molting, particularly among smaller crabs (< 100 g), which gained over 150% in total weight. Seasonal patterns were detected, with greater somatic gains in warmer months. No significant differences were observed between sexes overall, although some interactions were noted in specific morphometric traits. These findings suggest the potential occurrence of hypertrophic tissue expansion or mineral recycling in the absence of ecdysis, which merits further investigation. This study provides the first long-term physiological baseline for C. guanhumi under ex-situ conditions and supports its potential use in conservation and aquaculture programs.
Keywords: Terrestrial crustaceans; Ex-situ conservation; Reproductive biology; Captive management
Resumen
Este estudio evaluó la supervivencia, el crecimiento y el desempeño reproductivo de Cardisoma guanhumi durante 12 meses en condiciones semicontroladas. Se recolectaron 341 cangrejos adultos en manglares y se mantuvieron en un recinto naturalizado con refugios de PVC enterrados, estanques de agua dulce y salada, y rociadores automatizados para mantener la humedad. Los organismos fueron alimentados con una dieta vegetal suplementada con músculo de tilapia y monitoreados mensualmente. La supervivencia final fue de 76 y 68.6% de las hembras se volvieron ovígeras, evidenciando viabilidad reproductiva en cautiverio. Se registró crecimiento morfométrico en 170 individuos con apéndices intactos; el ancho del caparazón aumentó aproximadamente 1 mm por mes, pese a la casi ausencia de mudas visibles. Los organismos más pequeños (< 100 g) incrementaron su peso en más de 150%. Se observaron patrones estacionales, con mayores ganancias somáticas en meses cálidos. No se detectaron diferencias significativas entre sexos, salvo algunas interacciones en rasgos específicos. Los resultados sugieren posibles mecanismos de expansión tisular o reciclaje mineral sin muda. Este estudio aporta una línea base fisiológica a largo plazo y respalda el potencial de la especie para conservación y acuicultura.
Palabras clave: Crustáceos terrestres; Conservación ex-situ; Biología reproductiva; Manejo en cautiverio
Introduction
Coastal ecosystems provide critical services for biodiversity, food production, and human livelihoods. However, increasing anthropogenic pressures —including urban expansion, agriculture, and coastal infrastructure— have led to widespread loss and degradation of mangroves and wetlands (Alvarez & Briquets, 1983). This environmental decline directly threatens terrestrial crab species of the family Gecarcinidae, including Cardisoma guanhumi, a large land crab distributed from Florida to Brazil (Anger, 2013; Botelho et al., 2001).
Cardisoma guanhumi plays a key ecological role in mangrove ecosystems as a detritivore and sediment bioturbator. Its burrowing behavior enhances sediment oxygenation and stimulates microbial decomposition, contributing to nutrient cycling and substrate balance (Kristensen, 2008). Economically, it is highly valued in several Caribbean countries for its large chelipeds and meat content (Watson-Zink, 2021). In Mexico, although C. guanhumi is not currently listed under NOM-059 as endangered (Semarnat, 2020), its inclusion in local management programs reflects growing concern over population decline, and has been subject to seasonal fishing bans in regions such as Veracruz (Semarnat, 2020). Despite its ecological and commercial relevance, scientific knowledge on C. guanhumi is still limited, especially under captive conditions. Most available studies on C. guanhumi are over 2 decades old and focus primarily on natural history and reproductive cycles, with very limited data on captive performance or aquaculture potential (Diele et al., 2005; Smith & Smith, 2001).
Given the ongoing degradation of coastal habitats and the absence of recent biological data, understanding the growth and reproductive performance of C. guanhumi in captivity is essential. This study aimed to evaluate the feasibility of maintaining this species under semi-controlled conditions, describing seasonal growth patterns and reproductive output. The findings provide foundational data for conservation strategies and offer the first long-term evidence of physiological plasticity in captivity, supporting future aquaculture development for this native land crab. Such species are vital for coastal productivity and food web integrity (Gifford, 1963).
Materials and methods
The experiment was conducted at the Aquaculture Genetics and Production Laboratory of the Instituto Tecnológico de Boca del Río (ITBOCA), Veracruz, Mexico. A total of 341 adult specimens of C. guanhumi (132 males and 209 females; cephalothorax width > 50 mm) were manually collected during July and August 2022 from mangrove habitats along the Boca del Río-Antón Lizardo-Alvarado corridor (19°03’08” N, 96°00’27” W). Collection was performed under environmental authorization in compliance with national regulations and with minimal handling stress. Ovigerous females were excluded to avoid interference with growth and reproductive monitoring. Crabs were weighed (165.11 ± 50.19 g) and measured (CW: 6.80 ± 0.81 cm; CL: 5.06 ± 0.79 cm) before transport in 60-L plastic containers (15 individuals per container) to the laboratory facilities.
Crabs were maintained for 12 months in a semi-controlled, custom-built habitat designed to simulate key structural and microclimatic features of their natural environment. The enclosure measured 7 × 6 m, delimited by concrete perimeter walls 1.2 m high, and was installed directly over natural soil. A 30-cm-deep layer of fine sand was distributed across the entire area to allow natural burrowing behavior.
Eighty cylindrical PVC shelters, ranging from 2.5” to 6” in diameter, were buried diagonally into the substrate. Each tube was partially exposed at one end and fully embedded at the other, mimicking natural burrow structures and providing refuge. These shelters were distributed evenly across the enclosure. A young red mangrove (Rhizophora sp.) tree was planted at the center of the habitat to provide shading and promote environmental heterogeneity.
Two shallow basins were integrated at ground level to simulate aquatic access points: one freshwater basin (0.80 × 0.40 m) and one saltwater basin (0.60 × 0.50 m). These allowed the crabs to access both water types to mimic the species’ natural osmoregulatory behavior. The enclosure was protected with plastic mesh netting to prevent escapes and exclude predators.
A sprinkler system powered by a ½ HP submersible pump was installed to automatically maintain ambient humidity at 70 ± 10% and simulate rainfall events. This system operated intermittently throughout the day, creating a microclimatic regime resembling that of natural coastal mangroves. Environmental parameters including temperature and salinity were monitored daily using a YSI 556 multiparameter probe. Crabs were stocked at low density (< 1.0 ind/m²) to minimize territorial stress and promote natural behavior. A 5-day acclimation period was provided before experimental measurements commenced. The layout of the semi-controlled habitat and individual crab marking strategy are illustrated (Fig. 1).
Crabs were fed ad libitum with a rotating mixture of fruits and vegetables, supplied daily. Each individual received approximately 30-40 g/day of plant matter, including chopped papaya (10 g), mango (8 g), banana (5 g), coconut (3 g), watermelon (5 g), pumpkin (5 g), radish (2 g), and carrot (2 g), provided in alternating combinations to ensure dietary variety and stimulate natural foraging behavior.
In addition, each crab was offered 5 g of fresh tilapia (Oreochromis sp.) muscle every 2 days as a source of animal protein. All food items were weighed prior to distribution using a digital scale, and uneaten remnants were removed within 24 hours to prevent microbial proliferation and maintain substrate hygiene. Feeding was performed in the early morning to take advantage of natural crab activity rhythms. Food was distributed uniformly across the enclosure to reduce competition and ensure equal access among individuals.
A total of 170 crabs that retained all appendages throughout the experiment were selected for monthly morphometric monitoring. Each individual was permanently identified by engraving a unique mark on the dorsal carapace using a Dremel® rotary tool, a method shown to be effective and non-lethal for individual identification in crustaceans, followed by a dot of red nail polish to facilitate visual recognition during handling.
Nine morphometric parameters were recorded: total weight (TW), cephalothorax width (CW), cephalothorax length (CL), and bilateral measurements of the chelipeds, including: chelae (RCh, LCh), carpi (RCa, LCa), and meri (RM, LM). All linear dimensions were measured to the nearest 0.01 cm using a digital caliper (Autotec™), and weight was recorded to the nearest 2 g using an Ohaus Scout Pro SPU2001 electronic scale. Measurements were conducted at 30-day intervals for a total of 12 months. Handling was performed with care to minimize stress and avoid damage to limbs or exoskeleton. Data were logged immediately into individual digital records for subsequent statistical analysis. All measurements were performed by a single trained observer to reduce operator error.
Survival was assessed as the proportion of live individuals at the end of the 12-month experimental period relative to the initial number of crabs introduced into the habitat (n = 341). Daily inspections were conducted to detect mortalities, which were immediately removed to prevent deterioration and potential contamination of the enclosure. Causes of mortality were not determined but monitored for frequency to identify potential trends.
Reproductive activity was evaluated based on the presence of ovigerous females. Identification was performed by visual inspection of the ventral abdomen to detect external egg masses attached to the pleopods. Females were not handled unnecessarily to avoid detachment or disturbance of egg clutches. The number of ovigerous females was recorded seasonally, and reproductive peaks were determined based on their frequency during each sampling period. Although no copulation events were directly observed, the appearance of ovigerous females was interpreted as evidence of successful mating or possible sperm storage, as reported in other decapods. No interventions (e.g., hormone induction) were performed to stimulate reproduction.
Statistical analysis
Only crabs that retained all appendages throughout the entire experimental period (n = 170) were included in the statistical analyses. Data were first tested for normality using the Kolmogorov-Smirnov and Lilliefors tests, and for homogeneity of variances using Levene’s test. Sex-related differences across all morphometric variables were analyzed using one-way multivariate analysis of variance (MANOVA). To evaluate the effects of season and sex on specific traits, a two-way ANOVA was conducted for each morphometric variable independently. When significant differences were detected, Tukey’s HSD post hoc test was applied to identify specific group differences. Additionally, to assess growth performance by initial body size, crabs were grouped into 4 weight classes (0-99 g, 100-199 g, 200-299 g, > 300 g), and a two-way ANOVA (sex × weight class) was performed. Statistical analyses were performed using Statistica version 7.0 (StatSoft Inc., Tulsa, OK, USA), and differences were considered statistically significant at p < 0.05. Statistical procedures followed standard biometric methodology (Sokal & Rohlf, 2003).
Results
Cardisoma guanhumi was successfully maintained for 12 consecutive months in the semi-controlled habitat, showing a final survival rate of 76.25%. Males exhibited slightly higher survival than females. These values are comparable to those reported in similar experimental setups for other land or semi-terrestrial brachyurans, such as Gecarcinus lateralis and Ucides cordatus, where survival ranged between 70-85% under laboratory or mesocosm conditions (Greenaway, 2003; Hartnoll, 1982). However, published data on long-term captive maintenance of C. guanhumi remain scarce, with most available studies limited to short observational trials or reproductive monitoring under field conditions (Gifford, 1963; Shinozaki-Mendes et al., 2013). Sex-specific survival percentages are presented in Table 1.
Ovigerous females were recorded in all seasons, with the highest occurrence observed during summer and early autumn. A total of 68.6% of the females became ovigerous at least once throughout the year, consistent with historical accounts of reproductive periodicity in this species. Although no copulation events were directly observed during the trial, the consistent appearance of externally egg-bearing females suggests that successful mating occurred prior to or during the initial weeks of captivity. This observation supports the hypothesis of sperm storage capacity in C. guanhumi, a reproductive trait widely documented in decapod crustaceans and associated with multiple spawning events and temporal reproductive flexibility (Bauer, 1986; Taissoun, 1974).
On average, C. guanhumi exhibited a monthly morphometric gain of approximately 1 mm in cephalothorax width under captive conditions, despite the absence of molting in all but one individual. All morphometric variables showed significant seasonal increases over the 12-month period. The most pronounced gains occurred in cephalothorax width (CW) and total weight (TW), with maximum values recorded during spring and summer. These seasonal trends are illustrated in Figure 2. Two-way ANOVA indicated a significant effect of weight class on TW gain (p < 0.001), but no effect of sex (p > 0.05) nor sex × weight class interaction. Crabs in the lowest weight class (0-99 g) showed the highest relative TW gain (> 150%), whereas those above 300 g gained less than 3% (Fig. 2). Multivariate analysis of variance (MANOVA) did not detect overall differences between sexes across the 9 morphometric variables. However, a significant interaction was found for cephalothorax length (CL) between sex and season (p = 0.02).
Discussion
The relatively high survival rate can be attributed to the design of the artificial habitat, which included multiple buried shelters, regulated humidity through simulated rainfall, access to both freshwater and saltwater, and low stocking density. These conditions have been recognized as critical factors for terrestrial crab survival under controlled environments (Herreid, 1963; Oliveira et al., 2020), likely contributing to reduced stress, enhanced shelter availability, and overall behavioral stability.
Importantly, no signs of cannibalism or excessive agonistic behavior were observed during the experimental period. This contrasts with other decapods in captivity where confinement and competition for space often lead to significant mortality, as reported by Bauer (1986) and Fox (1975). The lack of mortality peaks during the reproductive season also suggests that the species can tolerate environmental manipulation and semi-controlled conditions without major physiological stress.
To our knowledge, this is one of the first experimental reports documenting long-term survival of C. guanhumi under captive conditions with ecological elements resembling natural mangrove habitats. Previous studies on this species have primarily been limited to short-term reproductive monitoring or natural history observations (Hartnoll, 1982; Smith & Smith, 2001).
Table 1
Summary of survival by sex in Cardisoma guanhumi after 12 months under semi-controlled conditions.
| Sex | Initial number | Final number | Survival (%) |
| Male | 132 | 106 | 80.3 |
| Female | 209 | 154 | 73.7 |
| Combined | 341 | 260 | 76.25 |
No statistical comparisons were performed, as crabs were maintained in a single habitat. However, survival rates by sex are reported for descriptive purposes.
The reproductive performance observed under semi-controlled conditions confirms that C. guanhumi is capable of adapting its reproductive cycle to artificial environments with minimal intervention. The presence of mangrove vegetation, natural soil substrate, and controlled humidity may have played a key role in maintaining environmental cues necessary for ovarian development and spawning behavior, as shown in studies on similar species (Mariappan et al., 2000; Silva & Oshiro, 2002). Unlike studies performed under purely laboratory conditions, where reproductive inhibition is common in terrestrial crabs, this setup offered a more ecologically relevant habitat, favoring reproductive success. The relatively high frequency of ovigerous females obtained without hormonal induction or photoperiod manipulation highlights the potential of C. guanhumi for captive broodstock development. Given the limited number of experimental studies on reproductive ecology in this species, these findings contribute novel evidence for its suitability in aquaculture and conservation-based culture systems.
The observed growth pattern in C. guanhumi under captive conditions may reflect a typical crustacean molting mechanism, in which the inner layer of the old exoskeleton is reabsorbed prior to ecdysis and a new cuticle is secreted and later calcified. This process allows size increase while maintaining structural stability of the carapace in adults. The reabsorption of calcium and other minerals, and their temporary storage in gastroliths or the hepatopancreas, has been well documented in decapods (Chang, 1995; Greenaway, 1985; Skinner, 1985), and is particularly relevant in land crabs due to limited environmental calcium availability.
Although somatic growth in decapod crustaceans is typically associated with molting events, the sustained morphometric increases observed in this study, in the apparent absence of clearly detectable ecdysis, may reflect subtle molting events, mineral recycling dynamics, or cuticular remodeling processes not externally observed. Similar mechanisms have been discussed in other decapods under stable environmental conditions (Chang, 1995; Greenaway, 1985; Skinner, 1985).
This physiological flexibility may represent an adaptive response of semi-terrestrial crabs to fluctuating environmental constraints (Anger, 2013; Zimmer et al., 2019). Further studies incorporating histological and endocrine approaches would help clarify the mechanisms underlying the morphometric patterns recorded here (Hartnoll, 2001). Similar sex-based morphometric assessments have been conducted in other brachyurans such as Menippe mercenaria, where cheliped development is linked to territorial behavior and reproductive strategies (Bauer, 1986; Savage & Sullivan, 1978).
Cardisoma guanhumi can be maintained under semi-controlled conditions with high survival and reproductive performance. The sustained morphometric growth recorded during the study provides a baseline for understanding its physiological responses in captivity and supports its potential for ex situ conservation and aquaculture applications.
Acknowledgments
The authors are grateful to Andrés Cabrera Muñoz from the Aquaculture Genetics and Production Laboratory (ITBOCA) for his technical assistance throughout the study.
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Wild chia (Salvia hispanica) populations, endangered under global warming scenarios
Peligro de extinción de poblaciones silvestres de chía (Salvia hispanica) bajo escenarios de calentamiento global
Sabina Lara-Cabrera a, *, Gleisery Rivas-Jaimes a, David A. Prieto-Torres b, c, Cuauhtémoc Sáenz-Romero c, d, Lourdes Núñez-Landa a, Guillermo Orozco-de Rosas e, Juan-Carlos Montero-Castro a
a Universidad Michoacana de San Nicolás de Hidalgo, Facultad de Biología, Laboratorio de Biodiversidad y Biogeografía de Plantas, Gral. Francisco J. Múgica s/n, Ciudad Universitaria, Edificio B2, 3er piso, Felicitas de Río, 58030 Morelia, Michoacán, Mexico
b Universidad Nacional Autónoma de México, Facultad de Estudios Superiores Iztacala, Laboratorio de Biodiversidad y Cambio Climático, Av. de los Barrios 1, Los Reyes Iztacala, 54090 Tlalnepantla, Estado de México, Mexico
c Laboratorio Nacional de Biología del Cambio Climático, SECIHTI, Mexico
d Universidad Michoacana de San Nicolás de Hidalgo, Instituto de Investigaciones sobre los Recursos Naturales, Av. Juanito Itzícuaro s/n, Col. Nueva Esperanza, 58337 Morelia, Michoacán, Mexico
e Chia Blanca, S.C. de R.L., La Paz No. 54, 45470 Acatic, Jalisco, Mexico
*Corresponding author: sabina.lara@umich.mx (S. Lara-Cabrera)
Received: 25 July 2025; accepted: 14 January 2026
Abstract
The conservation of crop wild relatives is particularly worrisome under global warming scenarios given their potential as sources of diversity for cultivars. Here we evaluate the potential effect of climate change through ecological niche modeling for wild chia (Salvia hispanica) populations as the closest crop wild relative. Current and future (2040, 2060, 2080) climatic conditions were modeled using 4 global climate models (ACCESS-CM2, BCC-CSM2-MR, MIROC6, and CanESM5) and 2 shared socioeconomic pathways (SSP2-4.5 and SSP5-8.5), accounting for both limited and long-distance dispersal scenarios. Current potential wild S. hispanica distribution is 192,789 km², 16.72% in climatically stable areas and coinciding with 184 Nature Conservancy Areas in Mexico and Guatemala. Unfortunately, by 2040, 2060, and 2080 wild chia’s potential distribution would reduce by 60.81 to 83.07%, respectively under optimistic scenarios with potential of dispersion as well as pessimistic and nondispersal scenarios. Furthermore, the most sensitive areas are in the Trans-Mexican Volcanic Belt, Pacific Coast, and Guatemala, which also are reportedly the most genetically diverse populations. We urge conservation efforts to increase wild germplasm collections and embark on in situ conservation with local communities to preserve the genetic reservoir of the species for future crop breeding.
Keywords: Climate change; Environmental suitability; Species distribution; Vulnerability risk
Resumen
Una de las mayores preocupaciones ante escenarios de cambio climático es la conservación de parientes silvestres de cultivos, dado su potencial como fuente de diversidad genética. Aquí evaluamos el efecto potencial del cambio climático actual y futuro (2040, 2060, 2080) para poblaciones silvestres de chía (Salvia hispanica), el pariente silvestre más cercano, a través de modelación del nicho ecológico utilizando 4 modelos de cambio climático (ACCESS-CM2, BCC-CSM2-MR, MIROC6 y CanESM5) bajo 2 trayectorias socioeconómicas (SSP2-4.5 y SSP58.5), considerando también dispersión limitada y de larga distancia. La distribución potencial actual de poblaciones silvestres de S. hispanica es de 192,789 km² y 16.72% en áreas climáticamente estables que coinciden con 184 áreas naturales de conservación de México y Guatemala. Desafortunadamente hacia 2040, 2060 y 2080 su distribución se reducirá de -60.81% a -83.70% respectivamente bajo trayectoria socioeconómica optimista con dispersión a larga distancia y pesimista con dispersión limitada. Adicionalmente las regiones más sensibles coinciden con las que se han reportado como genéticamente diversas en el Cinturón Volcánico Trans-Mexicano, Costa del Pacífico y Guatemala. Llamamos a incrementar los esfuerzos de conservación, aumentar las colecciones de germoplasma e iniciar proyectos de conservación in situ comunitarios con miras a conservar el reservorio genético del cultivo.
Palabras clave: Cambio climático; Idoneidad ambiental; Distribución de especies; Riesgo por vulnerabilidad
Introduction
Chia is a native crop of Mesoamerica (Cahill, 2003; Kirchhoff, 2000). Several attributes related to the high omega-3 content of chia (Salvia hispanica L.) have led to its steadily growing global consumption (Ali et al., 2012; Grancieri et al., 2019; Katunzi-Kilewela et al., 2021). While current estimates of global production increased in Paraguay, Bolivia, and Mexico from ca. 13,000,000 in 2019 to 18,000,000 in 2023 (Anuario Estadístico de Producción Agrícola, 2025) and consumption was estimated at 80,000 to 100,000 tons per year (Góral, 2025), little attention has been given to the distribution and conservation of wild chia populations, which are considered a Crop Wild Relative (CWR) and thus an important source of genetic variability for the crop (Maxted et al., 2006). The earliest documentary records, including the “Matrícula de tributos” (2022) and the “Relaciones geográficas de la Nueva España” (Murrieta-Flores et al., 2020), highlight the significance of chia in pre-Hispanic diets, ranking third in importance after maize and beans. However, the timeline of its domestication remains uncertain. The only unequivocal archaeobotanical evidence is nutlets dated to 1,750 years before present. Some studies suggest an earlier domestication event around 4,500 years ago based on pollen records (Sosa et al., 2016), although these may only be assignable to the genus Salvia rather than specifically to S. hispanica (LaraCabrera et al., 2025). Therefore, major questions regarding the identity and timeline of chia’s domestication cannot yet be addressed. Available evidence strongly suggests that domestication occurred in Mexico-Guatemala (i.e., Mesoamerica), a recognized center of origin for many globally important crops (Colunga-García Marín & Zizumbo-Villareal, 2004; Flannery, 1973). Chia has been included in Mexico’s inventory of 310 priority CWR species (Contreras-Toledo et al., 2018) and the global CWR inventory proposed by Harlan and De Wet (Vincent et al., 2013). The closest chia relatives are wild chia populations, making their study and conservation a pressing need. The conservation of CWRs is a global concern, as emphasized by the FAO, which recognizes their critical role in ensuring future food security in the context of global change (Kaeslin et al., 2012). In this context, the first step toward developing effective conservation strategies, as highlighted by Goettsch et al. (2021), is to generate accurate distribution maps, beginning by data curation to remove misidentified records and cultivated specimens. Here we apply ecological niche modeling using Maxent to wild chia populations in order to: 1) approximate the species’ Grinnellian niches (Peterson et al., 2011; Rödder & Engler, 2011); and 2) project future distribution under global warming scenarios (Dormann et al., 2007). Previous work by Durán et al. (2016) modeled environmental suitability for the periods 2041-2060 and 2061-2080, also using herbarium specimens. However, it remains unclear whether they distinguished wild from cultivated specimens, a key factor, as a mixed dataset could affect model accuracy due to differences in abiotic tolerance and dispersal capability. Here we carefully selected only wild chia specimens based on morphological traits, avoiding individuals exhibiting a domestication syndrome. For example, in wild populations, the fruiting calyx remains open, allowing nutlet release whereas in cultivated chia nutlets remain enclosed in the fruiting calyx. Chia nutlets (often misinterpreted as seeds) produce mucilage upon contact with water, and as has been described for thyme, the hydrated mericarp swells and pushes the calyx open to expose the nutlet (Bouman & Meeuse, 1992). Domesticated plants also tend to have larger overall size, inflorescences, and flowers (Cahill, 2003, 2005). Our dataset was subsequently used to model future niche distributions for the years 2040, 2060, and 2080 under optimistic and pessimistic Shared Socioeconomic Pathways (SSPs) following the Intergovernmental Panel on Climate Change (IPCC, 2022), global warming being one of the most severe threats to CWR conservation. We also considered chia’s potential for dispersal and colonization of new areas. Chia likely has limited natural dispersal via water (hydrochory or ombrochory; Zona, 2017), and when calices remain open gravity plays a role as nutlets fall near the parent plant and germinate nearby (phylomatri sensu Cheplick, 2021). Yet long-distance dispersal, potentially through human activity, may become crucial under accelerated climate change. Finally, given the serious conservation threat for many CWRs through land-use change (Goettsch et al., 2021), we evaluated the spatial overlap between the modeled distributions and the Mexican and Guatemalan Natural Conservation Areas. These areas represent an immediate opportunity for passive CWR conservation, especially at biodiversity hotspots where indigenous communities have managed the surrounding species for millennia resulting through domestication in a mosaic of landraces (Vincent et al., 2022). Our results offer valuable guidance for policymakers and plant breeders, enabling the identification of wild chia populations for in situ conservation and seed bank enhancement for ex situ conservation. Early, informed decisions are essential for preserving the genetic diversity of this culturally and nutritionally important crop.
Materials and methods
Biological data acquisition. Occurrence data follow Lara-Cabrera et al. (2025) identifying wild specimens from digitized herbarium collections at the Consortium of California Herbaria ([https://cch2.org/portal/collections/] RSA, UCR) MEXU (https://datosabiertos.unam.mx/ biodiversidad/), Portal Torch Herbaria ([https://portal. torcherbaria.org/portal/index.php] ASU, CIIDIR, COLO, DES, IBUG, IND, MICH, UNM, UTEP, WIS), US (https:// collections.nmnh.si.edu/search/botany/), and herbarium review (IEB). Domestication status was determined based on botanical descriptions (Klitgaard, 2012; Ramamoorthy, 1996; Wood & Harley, 1989) and cross-checked with studies on domestication syndrome traits in the species (Cahill, 2003, 2005; Calderón-Ruíz et al., 2021). Overall, wild specimens are recognized as shorter plants (30-100 cm tall), with short inflorescences (up to 10 cm), bearing 3 to 6 flowers per verticillaster, corolla tube 8-9 mm long that is almost entirely enclosed by the calix, and smaller fruiting calyces remaining open at maturity. Note that the ability to maintain open calyces to release the nutlets is a key trait distinguishing most domesticated varieties, as in the latter the calyx remains closed. All records were carefully reviewed to eliminate ambiguities such as duplicates, points located on roads or residential areas, and records with a spatial separation of less than 2 km to reduce spatial autocorrelation (Boria et al., 2014; Legendre, 1993; Peterson et al., 2011). Following Prieto-Torres et al. (2020, 2021) and PrietoTorres (2024), recent records (2001-2022) were subjected to an environmental outlier exclusion procedure. Records were removed if their values for Worldclim bioclimatic variables Bio 1 (annual mean temperature), Bio 5 (maximum temperature warmest month), Bio 12 (annual precipitation), and Bio 15 (precipitation seasonality) fell outside the upper or lower quartiles of the same variables calculated from herbarium records dated between 19702000. Those variables were used because other studies (e.g., Núñez-Landa et al., 2023; Prieto-Torres et al., 2020, 2021) have found them useful in the data depuration process. The final occurrence dataset included 55 independent localities of species presence, corresponding to the full list published by Lara-Cabrera et al. (2025). Environmental information for current and future scenarios. Available climatic layers were downloaded from Worldclim 2.1 (Fick & Hijmans, 2017; Hijmans et al., 2005) at a spatial resolution of 30 arc-seconds (~ 1 km²). These bioclimatic variables, derived from monthly temperature and precipitation data, reflect longterm climatic trends. Nonetheless, variables Bio 8 (mean temperature wettest quarter), Bio 9 (mean temperature dries quarter), Bio 18 (precipitation of wettest quarter), and Bio 19 (precipitation of coldest quarter) were excluded due to potential spatial artifacts in their interpolated surfaces, which could negatively affect model performance (Booth, 2022; Escobar et al., 2014). Checking for such anomalies is particularly important when modeling across broad regions, as in this study, to prevent localized irregularities which could bias overall results (Booth, 2022). And although soil properties influence species distributions by affecting plant physiology, growth, and survival (Eamus et al., 2013; Velazco et al., 2017), we focused exclusively on climatic variables because no reliable projections of edaphic variables exist for future scenarios (Tomlinson et al., 2020). Moreover, temperature and precipitation strongly influence soil properties, so climatic variables are expected to indirectly capture part of the soil-related variability. This reasoning is consistent with studies highlighting climate as the main driver for long-term distributional shifts (Dantas et al., 2020; Hartmann et al., 2022). To avoid model overfitting due to collinearity among variables (Dormann et al., 2013), we employed a twostep process to select relevant predictors and reduce dimensionality (Cobos et al., 2019). First, Spearman and Pearson correlation coefficients were used to exclude variables with correlations greater than 0.8 and with a variance inflation value greater than 4 (Díaz-Vallejo et al., 2024; Dupin & Smith, 2019). Both coefficients were used because normality tests indicated that each variable could exhibit different spatial patterns, and a single test would not be feasible due to high spatial autocorrelation among variables, thereby justifying the combined use of both correlation measures (Díaz-Vallejo et al., 2024). Second, a principal component analysis (PCA, Janeković & Novak, 2012) was performed on the original 15 climatic variables across the study area to retain those that collectively explained at least 95% of the total variance (Hanspach et al., 2011). These procedures were performed in R using libraries “corplot”; “usdm” (Wei & Simko, 2017), and “ENMGadgets” (Barve & Barve, 2016). The best climatic data approach was selected based on statistical performance evaluated with the “kuenm” R package (Cobos et al., 2019). For future climate projections, we considered 2 Shared Socioeconomic Pathways from the Coupled Model Intercomparison Project (CMIP6): SSP2-4.5, representing a middle-of-the-road scenario with moderate emissions; and SSP5-8.5, a high-emission scenario reflecting continued fossil fuel dependence and rapid economic growth (Riahi et al., 2017). Projections were generated for 3 future timeframes: 2041–2060 (hereafter 2040), 2061– 2080 (hereafter 2060), and 2070-2090 (hereafter 2080), using 4 General Circulation Models: ACCESS-CM2, BCCCSM2-MR, MIROC6, and CanESM5. These models were selected based on their ability to simulate ENSO dynamics and associated interannual variability in precipitation and temperature (Bi et al., 2013; Watanabe et al., 2010; Zelinka et al., 2020), which are key drivers of climate patterns within the range of S. hispanica and and the whole biota of Central American and Mexico. Therefore, this selection ensures that the models capture climate processes directly relevant to the species’ environmental niche and potential distribution (Jin et al., 2025). The environmental variables selected were cropped to a region defined as M used for model calibration processes (Barve et al., 2011; Soberón & Peterson, 2005). The vectorial polygon defining M was constructed by intersecting species occurrence records with maps of terrestrial ecoregions (Dinerstein et al., 2017) and Neotropical Biogeographic provinces (Morrone et al., 2017). Although several valid approaches exist to define the M area in ecological niche modeling, we selected this method because it is not only straightforward but also highly operational (Rojas-Soto et al., 2024). Ecological niche and potential species distribution. Species models were performed with Maxent 3.4.3 (Phillips et al., 2006) and the R package “kuenm” (Cobos et al., 2019). Maxent constructs the most probable species distribution based on presence-only data and environmental variables through machine learning algorithms, as it is recognized as one of the most effective tools in terms of predictive performance (Elith et al., 2006). The “kuenm” R package was used to generate multiple candidate models by combining different sets of environmental variables (non-correlated and PCAreduced) with a range of Maxent calibration parameters (described below) to select the optimal model and subsequently project it to climate scenarios for 2040, 2060, and 2080. Occurrence records were initially split into 2 datasets: 80% were randomly selected for model training, while the remaining 20% were used for internal validation of the final model. During the calibration step, we created spatial folds using 80% of the records for training, and spatial cross-validation was applied. This approach allowed us to test 992 candidate models resulting from all 62 combinations of Maxent feature classes and 8 regularization multipliers (0.2, 0.4, 0.5, 1, 2, 6, 8, 10) and the 2 sets of variables (non-correlated and PCA-derived), in order to select the optimal set for the final modeling (Cobos et al., 2019). Optimal models were selected based on 3 criteria: statistical significance via partial ROC partial test (Peterson et al., 2008), Akaike’s information criterion for small sample sizes (AICc; Akaike, 1974), and an omission rate below 5% (Anderson et al., 2003). Final models were generated using 1,000 iterations and 10 replicates. Next, the models were projected onto future climate scenarios using 3 extrapolation settings: no extrapolation or clamping, extrapolation without clamping, and both extrapolation and clamping. These approaches are essential to identify novel climatic conditions under future scenarios, potentially suitable for the species based on extreme values of ecological variables (Elith et al., 2011; Peterson et al., 2018; Stohlgren et al., 2001). In this context, to detect future regions with non-analogous climatic conditions relative to the present, we implemented the MOP (Mobility-Oriented Parity) test (Owens et al., 2013). This test identifies zones that represent strict extrapolations of the model, which are associated with higher uncertainty and should be interpreted with caution (Alkishe et al., 2017). The Maxent outputs in “cloglog” format were converted to binary presence-absence maps using a threshold equal to the 10th percentile of training presence. This threshold minimizes commission errors by excluding only the lowest 10% of suitability scores, thus retaining predictions considered at least as suitable as the occurrence records (Liu et al., 2013). The binary maps for each Shared Socioeconomic Pathway (SSP2-4.5 and SSP5-8.5) and time period (2040, 2060, and 2080) were generated by overlaying the outputs from 4 selected Atmosphere-Ocean General Circulation Models (ACCES-CM2, BCC-CSM2MR, MIROC6, and CanESM5). Herein, a grid cell was classified as “presence” if at least 3 of the 4 models agreed on the presence prediction for that cell (e.g., Nuñez-Landa et al., 2023; Gama-Rodríguez et al., 2024). Spatial analyses and general metrics. The current binary map was compared with each future projection by overlaying and summing them, to define climatically stable areas, where a cell was considered “climatically stable” if the species’ environmental conditions were predicted therein for all evaluated periods (Collevatti et al., 2013). It is important to note that this does not imply that the local climate remains unchanged, but rather that the cell stays within the species’ tolerable range, allowing potential persistence (Terribile et al., 2012). Identifying such areas is important because they often harbor higher intraspecific genetic diversity (Hewitt, 2004). Besides, a range gain was defined when the number of pixels predicted to be suitable for distribution areas in the future were greater than those estimated in the present, and the opposite case was interpreted as range contraction for the species. Given that a species’ dispersal capacity can influence its ability to colonize new areas (Marco et al., 2011), we conducted this evaluation under 2 contrasting dispersal scenarios: short-distance dispersal (SDD), where suitable future areas were limited to those overlapping with the current distribution, and long-distance dispersal (LDD), where suitable areas projected for the future but not currently occupied were also considered. In both cases, geographic barriers defined by the M area were assumed to limit species dispersal, and ecological interactions between species were omitted (Atauchi et al., 2020; Peterson et al., 2002). For projections predicting a loss of suitable areas, we calculated the differences (current vs. future) in values of the bioclimatic and elevation variables to identify the environmental changes associated with this loss (Cobos & Bosch, 2018; Atauchi et al., 2020). The geographic centroid of the binary maps (for each present and future projections) was computed using the coordinates of all presence-value cells, with the constraint that the centroid must lay within the presence area. To enhance discussion, centroid shift was estimated in all scenarios and temporalities and actual km² and displacement direction estimated in R through the haversine formula proposed by Inman (1835). Additionally, the mean elevation and its range for each temporality were extracted by coupling future climate projection pixels with a digital elevation model. Both analyses were performed with the “terra” R package (Hijmans, 2025) Finally, we assessed the overlap between the current and projected species distribution and Natural Protected Areas (NPAs) in Mexico (https://www.gob. mx/conanp#1692) and Guatemala (https://conap.gob.gt/ listado-de-areas-protegidas/; IUCN & UNEP-WCMC, 2025). This was done by intersecting current and future binary distribution maps with the raster layers of official NPAs to identify key regions where in situ conservation of chia may be crucial in the face of land use change, contributing to risk mitigation through habitat protection.
Results
Current distribution pattern. The potential distribution model for S. hispanica was adequately calibrated, as indicated by the mean AUC ratio (1.85), minimum AIC value (708.46), statistically significant partial ROC test (0.00), and 0 omission error. The default plot produced shows both omission rates and AICc values for all tested models (Supplementary material: Fig. 1). The final model used a regularization multiplier of 2 and included the feature classes quadratic, product, and hinge. The most important variables that define the first principal component (PC1) are Bio1, Bio3, Bio6, Bio11, and Bio12, whereas the second component (PC2) is primarily defined by Bio12, Bio14, Bio15, and Bio17. The estimated current potential distribution area for wild chia is 192,789 km², primarily located across the biogeographic regions of the Sierra Madre Occidental, Trans-Mexican Volcanic Belt, Sierra Madre del Sur, and Sierra Madre de Chiapas, extending into Guatemala (Fig. 1). In Mexico, chia is distributed from the northeastern states (Sonora, Chihuahua, Durango, and Sinaloa), through the western and central regions, and into the southeast (Guerrero, Oaxaca, and Chiapas). To assess average population shift regarding latitude and longitude we estimated the centroid at 21.0367 oN, -102.231 oW (Jalisco Altos Sur, NW Jalisco, Mexico), average elevation 1,971.1 m and elevational range 1708.6 – 2211.3 m (Table 1). A valuable finding is that 11.51% (i.e., 22,206 km²) of the current projected distribution lies within NPAs. In Mexico the current potential distribution overlaps with
135 NPAs, including 9 Biosphere Reserves (UNESCOMAB), Flora and Fauna Protected Areas, Protected Areas for Natural Resources, National, State, Municipal Parks, and Voluntary Conservation Areas. Whereas in Guatemala the current potential distribution overlaps with 49 areas, including one Biosphere Reserve (UNESCOMAB), Natural Reserves, National and Regional Parks, Definitive Veda Zones, Forest Reserves and Use Reserves (Supplementary material: Table 1). Projected distribution patterns under non-dispersal and dispersal scenarios. The potential distribution of wild S. hispanica populations for 2040, 2060, and 2080, modeled considering non dispersal (i.e., SDD), highlights a concerning and pessimistic trend of habitat loss (Table 1, Fig. 2a). Under the optimistic global warming SSP24.5 scenario, the potential distribution area is projected to lose 38.27% by 2040, 50.07% by 2060 and 60.81% by 2080, leaving only a remaining area of 119,002 km², 96,250 km², and 75,551 km², respectively. In contrast, under the pessimistic warming scenario (SPS5-8.5), the remaining area is expected to be 32,638 km², reflecting a dramatic 83.07% loss of the current distribution. The loss of suitable habitat also extends to populations currently located within NPA: by 2040 only 15,879 km² (i.e., 13.34% of remaining suitability areas) would be within NPA and by 2080 these values are projected to decline further to 11,502 km² (i.e., 15.22%) under the optimistic scenario, and 6,195 km² (18.98%) under the pessimistic one. Furthermore, the geographic centroid of the species’ distribution is expected to shift in both longitude and latitude, accompanied by notable elevational shifts. In the SSP2-4.5 scenario, the mean elevation (currently 1,971.1 m) is projected to increase to 2,246.4 m (range: 2,0572,420.1 m), and to 2,376 m (range: 2,194-2,542.1 m) under SSP5-8.5. This corresponds to an average upward shift in elevation of 275.3 m and 405 m, respectively, along with a contraction of the current elevational range from 503 m to 357 m and 348 m under each scenario. Conversely, under the dispersal scenario (i.e., LDD), the projected future distributions are less alarming, although still indicative of significant habitat changes. These models assume the species eventually could colonize newly suitable areas, although likely requiring human-assisted dispersal. Projected distributions under
| Models | Potential distribution (pixels / km²) | Within NPA — Pixels (%) | % present vs. future | Change proportion attributable to GCC | Centroid displacement (km) and direction | Average elevation (m) | Min. elevation (m) | Max. elevation (m) |
|---|---|---|---|---|---|---|---|---|
| Current | 192,789 | 22,206 (11.51) | – | – | 21.0367, -102.231 | 1,971.1 | 1,708.6 | 2,211.3 |
| 2040 SSP2-4.5 | 119,002 | 15,879 (13.34) | 8.24 | -38.27 | 76.51 towards NW | 2,126.7 | 1,912.6 | 2,324.5 |
| 2060 SSP2-4.5 | 96,250 | 13,701 (14.23) | 7.1 | -50.07 | 81.13 towards NW | 2,184.9 | 1,984.1 | 2,372.2 |
| 2080 SSP2-4.5 | 75,551 | 11,502 (15.22) | 5.9 | -60.81 | 89.8 towards N-NW | 2,246.4 | 2,057 | 2,420.1 |
| 2040 SSP5-8.5 | 117,304 | 15,622 (13.32) | 8.1 | -39.15 | 14.71 towards SW | 2,130.4 | 1,915.5 | 2,329.6 |
| 2060 SSP5-8.5 | 77,794 | 12,060 (15.50) | 6.3 | -59.65 | 90.63 towards N-NW | 2,233 | 2,035.2 | 2,413.6 |
| 2080 SSP5-8.5 | 32,638 | 6,195 (18.98) | 3.2 | -83.07 | 84.18 towards N-NW | 2,376 | 2,194 | 2,542.1 |
| Models | Potential distribution (pixels / km²) | Within NPA — pixels (%) | % present vs. future | Change proportion attributable to GCC | New area — pixels / km² | New area — % | Centroid displacement (km) and direction | Average elevation (m) | Min. elevation (m) | Max. elevation (m) |
|---|---|---|---|---|---|---|---|---|---|---|
| Current | 192,789 | 22,206 (11.51) | – | – | – | – | 21.0367, -102.231 | 1,971.1 | 1,708.6 | 2,211.3 |
| 2040 SSP2-4.5 | 145,180 | 19,217 (13.24) | 9.96 | -24.69 | 26,178 | 18.03 | 114.4 towards NNW | 2,182.8 | 1,960.6 | 2,390.6 |
| 2060 SSP2-4.5 | 137,076 | 19,513 (14.24) | 10.12 | -28.90 | 40,826 | 29.78 | 140.85 towards NW | 2,277.5 | 2,065 | 2,475.4 |
| 2080 SSP2-4.5 | 125,364 | 18,474 (14.74) | 9.58 | -34.97 | 49,813 | 39.73 | 151.64 towards NW | 2,365.1 | 2,173.8 | 2,542.3 |
| 2040 SSP5-8.5 | 149,217 | 19,778 (13.25) | 10.25 | -22.60 | 31,913 | 21.39 | 117.54 towards NNW | 2,197.8 | 1,972.1 | 2,409.6 |
| 2060 SSP5-8.5 | 123,825 | 18,499 (14.94) | 9.59 | -35.77 | 46,031 | 37.17 | 127.45 towards NNW | 2,341.8 | 2,141 | 2,527.1 |
| 2080 SSP5-8.5 | 78,662 | 14,117 (17.95) | 7.3 | -59.20 | 46,024 | 37.17 | 202.01 towards NW | 2,508.4 | 2,346.2 | 2,655.3 |
Projected distributions under both SSP2-4.5 and SSP5-8.5 scenarios are relatively similar (Table 2, Fig. 2b). Under SSP2-4.5, the species is projected to lose 24.69% of its current range (remaining area: 145,180 km²), while gaining approximately 26,178 km² (18.03%) of new areas by 2040. However, by 2080, only 65% of suitable habitat will persist, with a loss of 35% of the area. Moreover, the proportion of the species’ range within NPAs is also projected to decline under dispersal scenarios, from current potential distribution to future: 18,474 km² (i.e., 9.58% of remaining areas) under SSP2-4.5 and 14,117 km² (i.e., 7.3%) under SSP5-8.5 by 2080. Projected centroid shifts (Table 2) are expected to be up to 114.4 km towards the N-NW and 151.64 km NW by 2080 (S-SW state of Zacatecas), with an optimal elevation rising to 2,365.1 m on average (range: 2,173.8–2,542.3 m) under the optimistic scenario. Under a pessimistic scenario the projected centroid would shift 117.54 km towards N-NW and by 2080 up to 202.01 km towards the NW. We estimated a shift of 2,508.4 m (range: 2,346.2–2,655.3 m), reflecting an upward elevational gain of 394 m to 537.3 m, and a narrowing range of the elevation from 503 m to 364 m and 309 m, respectively.
Long-term climatically stable areas and environmentally driven area loss. Modeled area loss by 2040 and 2080 under a pessimistic scenario was primarily driven by increases in temperature and changes in precipitation regimes (Table 3). Average annual temperature is projected to increase in 1.61 to 4.15 °C, with similar increases for the warmest month (+1.84 to +4.77 °C) and the coolest month (+1.49 to +3.78 °C). Annual and seasonal precipitation are expected to increase slightly (+33.97 to +9.77 mm; +2.10 to +2.64%) from current levels, while precipitation is projected to decrease in the driest quarter (0.92 to 3.18 mm). Under pessimistic scenarios, environmentally suitable areas for species are associated with elevated temperature values across multiple climatic variables, including average annual temperature (+1.83 to +4.17 °C), maximum temperature of the warmest month (+1.83 to +4.78 °C), and minimum temperature of the coldest month (+1.49 to +3.78 °C). The coefficient of variation for seasonal precipitation also increases moderately (+2.14 to +2.91%), while specific reductions are projected for the driest quarter (0.93 to 3.31 mm) compared to current conditions. This seasonal imbalance suggests harsher conditions during warm and dry periods, despite total precipitation increases.
Finally, the climatically stable area suitable for S. hispanica was estimated to be 32,245 km², representing only 16.72% of its total current distribution area (Fig. 3). In Mexico, these stable areas (representing 15.5% of total estimated sites) are located primarily in the northwest (states of Chihuahua and Durango), west (Jalisco and Michoacán), central south (Estado de México, Mexico City, and Guerrero), and southern regions (Oaxaca and Chiapas).
| Variable | SSP2-4.5 (Optimistic scenario) | SSP5-8.5 (Pessimistic scenario) | ||||
|---|---|---|---|---|---|---|
| 2040 | 2060 | 2080 | 2040 | 2060 | 2080 | |
| bio 1 Annual mean temperature (°C) | +1.61 | +2.71 | +4.15 | +1.61 | +2.71 | +4.17 |
| bio 2 Mean diurnal range (mean of monthly (max temp − min temp)) (°C) | +0.07 | +0.21 | +0.41 | +0.07 | +0.20 | +0.41 |
| bio 3 Isothermality (BIO2 / BIO7) (× 100) | -0.55 | -0.76 | -0.89 | -0.55 | -0.78 | -0.91 |
| bio 4 Temperature seasonality (standard deviation × 100) | -1.65 | +0.59 | +4.50 | -1.71 | +0.45 | +4.20 |
| bio 5 Max temperature of warmest month (°C) | +1.84 | +3.13 | +4.77 | +1.83 | +3.14 | +4.78 |
| bio 6 Min temperature of coldest month (°C) | +1.49 | +2.49 | +3.78 | +1.49 | +2.49 | +3.78 |
| bio 7 Temperature annual range (BIO5−BIO6) (°C) | +0.34 | +0.64 | +0.99 | +0.34 | +0.64 | +1.00 |
| bio 10 Mean temperature of warmest quarter (°C) | +1.58 | +2.73 | +4.20 | +1.58 | +2.72 | +4.20 |
| bio 11 Mean temperature of coldest quarter (°C) | +1.62 | +2.70 | +4.09 | +1.62 | +2.71 | +4.11 |
| bio 12 Annual precipitation (mm) | +33.97 | +27.62 | +9.77 | +34.76 | +29.44 | +11.39 |
| bio 13 Precipitation of wettest month (mm) | +16.74 | +16.39 | +12.97 | +16.93 | +16.88 | +14.36 |
| bio 14 Precipitation of driest month (mm) | -0.10 | -0.38 | -0.66 | -0.10 | -0.40 | -0.72 |
| bio 15 Precipitation seasonality (coefficient of variation) (%) | +2.10 | +2.41 | +2.64 | +2.14 | +2.53 | +2.91 |
| bio 16 Precipitation of wettest quarter (mm) | +22.52 | +18.74 | +2.34 | +23.28 | +20.44 | +4.86 |
| bio 17 Precipitation of driest quarter (mm) | -0.92 | -1.95 | -3.18 | -0.93 | -2.00 | -3.31 |
| Climatically stable areas in km² | 32,245 | |||||
| % | 16.72 | |||||
Discussion
Here we report a concerning scenario for the persistence of wild chia populations in the face of global warming through a Species Distribution pipeline employed in Maxtent. Although other pipelines are available such as the use of Random Forests (Breiman, 2001), we were unable to employ it because it requires an even number both of species presences and species absences, which is difficult since usually the latter is many folds larger than the number of presences (Gómez-Pineda et al., 2020). Nevertheless, our resulting hypotheses are robust following current scientific criteria and could serve as a starting point to guide conservation of this CWR for in field follow up to monitor wild chia populations, to reassess proposed abiotic variables and soil interacting agents to fine tune critical variables, and to enhance hypotheses and assure assisted migration success if implemented. Climate change poses imminent threats to individuals, species, and ecosystems (Kaeslin et al., 2012), and the conservation of CWRs is directly linked to future food security. The projected future distribution scenarios for 2040, 2060, and 2080 indicate a dramatic decline in the extent of suitable habitats for wild chia populations; even under the most optimistic scenario area loss is projected. We also foresee a geographic shift in the species’ average longitude and latitude climatic niche, along with a sharp net contraction in elevational range, where the lower elevation limit of the current natural range is expected to suffer larger losses; this finding agrees with projections for many other wild plant species through Mexican mountain gradients (Ramírez-Barahona et al., 2025). The speed of climatic change, exacerbated by the current acceleration of this process (Hansen et al., 2025), makes it highly unlikely that natural populations could disperse their seeds and establish new populations at the locations where suitable climatic conditions will emerge at the required pace. In other words, the mismatch between the areas currently occupied by wild populations and the shifting climate suitable for them will increase over time. Human-assisted translocation of seeds would likely be necessary to ensure species persistence. On the other hand, so-called “climatically stable” areas should not be interpreted with undue optimism. The persistence of an area within the species’ overall climatic niche does not imply that local conditions remain unchanged. This distinction is important when considering populations with significant genetic differentiation in traits such as frost resistance or drought tolerance. Specific populations may remain within the species’ broad climatic niche but become uncoupled from the narrow climate interval to which they have evolved and specialized, as demonstrated for wild forest tree populations along environmental gradients (Ortiz-Bibian et al., 2017; Rehfeldt et al., 2018). All these results add S. hispanica to the growing list of CWRs at risk. For instance, Jarvis et al. (2008) projected climate change effects on the conservation of wild relatives of peanuts (Arachis), potato (Solanum), and cowpea (Vigna), emphasizing the urgency of preserving their natural habitats and, in extreme cases, collecting and storing their genetic material in germplasm banks. Alarmingly, wild populations in the Trans-Mexican Volcanic Belt, a region of highest reported genetic diversity for chia (Cahill, 2004), are expected to be completely lost. Our global climate change scenarios differ markedly from those reported by Durán et al. (2016), who projected an increase in the potential distribution of S. hispanica (chia) under future climate conditions. Overall, they interpreted global climate change as potentially beneficial for the species, predicting moderate to high environmental suitability for 2060 and 2080, and an expansion of suitable areas across central and southern Mexico (Durán et al., 2016). Direct comparison between both studies is difficult due to several methodological differences, including the source of occurrence records (wild-only vs. potentially mixed wild and cultivated), variation in raster resolution (2.5 arc-minutes vs. our 30 arc-seconds), and differences in environmental variables. While we relied solely on bioclimatic variables from WorldClim, Durán et al. (2016) supplemented these data with elevational information from SIAN (translated as Environmental Information System) at INIFAP (National Institute of Forestry, Agriculture and Livestock Research) and photoperiod. Unlike Durán et al. (2016), we did not include elevation. Elevation could be considered a surrogated variable of Mean Annual Temperatures and related temperature variables. There is a strong correlation between temperature and elevation along elevational gradients, at a lapse rate of approximately 0.5 oC for each 100 m of elevational difference (Sáenz-Romero et al., 2010). However, when the species has extensive latitudinal distribution (as in this case) the correlation temperature-elevation will vary and thus, adding elevation as an environmental variable could have confounded effects. Nonetheless, both studies converge in predicting habitat loss in northeastern Mexico, and project similar temperature increases: Durán et al. (2016) estimated an average temperature rise of 1 to 3 oC, while our models predict an average temperature increase of 2.28 oC, ranging from 1.61 to 4.16 oC over the average. Additionally, our projections indicate a mean annual precipitation increase, although it is extremely important to realize that such change has a strong seasonal component. Actual precipitation increase is projected only for the wettest quarter; meanwhile, there is a net decrease of precipitation in all the other quarters. Thus, the total annual precipitation increase is not necessarily beneficial for the biology of the species, since the dry season becomes both drier and warmer. The balance between temperature and rainfall variables are critical for the species’ survival and ecological performance (Böhning-Gaese et al., 2008; Uribe, 2015). Although our study focused exclusively on wild populations of chia as a CWR, it is relevant to note that projections for chia cultivation in Mexico by 2040-2069 (Orozco de Rosas et al., 2014) predict an expansion of suitable areas, particularly in the highlands of Sinaloa, Nuevo León, Tamaulipas, Veracruz, Nayarit, Jalisco, Michoacán, Guerrero, Oaxaca, Morelos, Puebla, and Chiapas. These findings do not contradict our results, as they reflect different environmental requirements between wild and cultivated chia (Chen et al., 2017). Unfortunately, if the CWRs are lost, new characters would not be available to transfer into the cultivars, as has been the case in other crops, such as resistance to pests [from Solanum demissum Lindl. (Díaz-García et al., 2023) to potatoes; from Solanum peruvianum L., Solanum cheesmanii (L.Riley) Fosberg, and Solanum pennellii Correll to tomatoes (Rick & Chetelat, 1995)]; cytoplasmic male sterility in sunflowers (Hajjar & Hodgkin, 2007); and tolerance to drought and salinity in tomatoes (Solanum chilense (Dunal) Reiche and S. pennellii (Rick & Chetelat, 1995) among other traits (see Hajjar & Hodgkin, 2007 for further information). To date, most breeding efforts in chia have focused on mutation induction via radiation (Jamboonsri et al., 2012; Sorondo, 2017), and the full potential of wild relatives remains largely unexplored. Although no severe pestrelated threats have yet been reported for chia cultivars, global warming is enabling the emergence of new pests in many crops (Srinivasa et al., 2022), and chia may face similar risks in the near future. Therefore, we need to be prepared. For example, in Ghana, where chia is cultivated to alleviate hunger in low-income communities due to its low production costs and high nutritional value, several pest infections have already been reported, including coreid bugs (Lagria sp.), Zonocerus variegatus L., Diopsis macrophthalma Dalman (reported as the synonym Diopsis thoracica Westwood) and Fusarium wilt (Yeboah et al., 2014). In Mexico, chia’s center of origin, stem damage by Diabrotica speciosa (Germar), 1824 (Coleoptera) has been shown to reduce seed yield by up to 43% (Sosa-Baldivia & Ruiz-Ibarra, 2016). Interestingly, Salvia karwinskii Benth., a taxonomically close wild relative of chia, was reported to exhibit strong antifeedant activity against Spodoptera littoralis (Boisduval) and Diabrotica virgifera virgifera Leconti (Esquivel et al., 1996), highlighting the value of conserving wild taxa. Beyond climate change, other threats must be addressed to ensure effective in situ conservation of wild chia populations. Among these, land-use change related to anthropogenic activities is particularly severe and rapidly accelerating (Davison et al., 2021; Goettsch et al., 2021; Villaseñor et al., 2024). Although we were unable to model this factor directly, chia’s habitat loss has been documented since 2003 in regions such as Huehuetenango (Guatemala) and in the Mexican states of Jalisco, Colima, and Michoacán (Cahill, 2003) mainly due to overgrazing, land-use transformation to agricultural or residential, herbicide application in nearby monocultures, and induced or accidental burning (personal observations GOR). The current rates of habitat loss and degradation are even more dramatic, particularly due to transformation of pineoak woodlands, which are rich in Salvia species -other CWRs- (Cornejo-Tenorio & Ibarra-Manríquez, 2011; Lara-Cabrera et al., 2016). A hopeful sign is the overlap between the projected distribution of wild chia and existing conservation areas in Mesoamerica. As with other CWRs, these protected areas could serve as strongholds for in situ conservation (Tobón-Niedfeld et al., 2022). And although CWRs are not commonly used to define protected areas, they have been employed in specific cases, as in the establishment of the Sierra de Manantlán Biosphere Preserve to protect the corn wild relative Zea diploperennis (Guzmán & Iltis, 1991). New protected areas could be established to preserve the ca. 100 Mesoamerican CWRs threathened with extinction (Goettsch et al., 2021). Therefore, integrating CWR monitoring into conservation area management is a valuable recommendation (Vincent et al., 2022). Fortunately, wild chia currently occurs, and is expected to persist, within some of the most emblematic Biosphere Reserves in Mexico and Guatemala. We recommend in situ conservation managed by local communities, who have cultivated and utilized chia for over 3,000 years; this should be complemented by ex situ strategies, such as establishing germplasm banks that ensure adequate sampling of wild populations. However, access to this material must be regulated to protect local interests assuming The International Treaty on Plant Genetic Resources for Food and Agriculture (https://www. fao.org/plant-treaty/overview/text-treaty/en). Also, ex situ conservation could involve cultivating wild populations in areas projected to remain suitable or those that would emerge as new suitable areas (in general at higher elevations than today) under future climate scenarios, as proposed for other rare and endangered Mexican plant species (Mendoza-Maya et al., 2022).
Acknowledgements
The first author acknowledges financing through the Secretaría de Ciencias, Humanidades, Tecnología e Innovación: Project 319466; Coordinación de la Investigación Científica, Universidad Michoacana de San Nicolás de Hidalgo. We also thank Joseph Cahill for access to his data set, Yocupitzia Ramírez Amezcua for fruitful discussions, Pedro Chamu for plant collections, and Geraldine Murillo and David Lomeli who participated in data set depuration.
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Strategies of space use and foraging effort in two Neotropical flocking warbler species (Parulidae) during their nonbreeding period
Estrategias del uso del espacio y esfuerzo de forrajeo en dos especies de chipe Neotropicales (Parulidae) de parvada durante su periodo no reproductivo
Emerenciano Rivera a, *, Brian McLaren a, Eric Hernández-Molina b
a Faculty of Natural Resources Management, Lakehead University, 955 Oliver Road, Thunder Bay, P7B5E1 Ontario, Canada
b Pronatura Sur, A.C., Calle Dr. Navarro No. 5, Barrio El Cerrillo, 29220 San Cristóbal de las Casas, Chiapas, Mexico
*Corresponding author: erivera@lakeheadu.ca (E. Rivera)
Received: 22 August 2025; accepted: 24 December 2025
Abstract
We explored habitat use and foraging effort in 2 migratory species, the Golden-cheeked warbler and the Townsend’s warbler (Parulidae) to identify wintering habitats and foraging costs associated to group behavior. First, we compared densities, sex/age ratios and use of foraging substrates within and between species as a function of habitat conditions in 2 landscapes of varying forest cover. Then, we estimated home range and proxies of foraging effort to identify potential fitness costs of group behavior. Densities along with sex and age ratios of the 2 focal species differed between higher and lower quality sites, with adult males of both species being more abundant in flocks. Higher frequency in use of inner branches and trunks and lower stratums such as the understory at the micro-site level was observed in the Townsend’s warbler, whereas use of outermost leaves, twigs and upper stratums such as canopy were observed in the Golden-cheeked warbler. Home ranges of individual Townsend’s warblers overlapped areas used by flocks suggesting potential benefits of retaining a territory in areas where flocks occur. Traveled distances, daily used areas and searching times were higher in lower-quality sites for both species with increased group size as an underlying variable influencing cumulative searching times.
Keywords: Competition; Golden-cheeked warbler; Mixed-species bird flocks; Neotropics
Resumen
Exploramos el uso de hábitat y esfuerzo de forrajeo en 2 especies migratorias, el chipe de cachetes amarillos y el chipe de Townsend (Parulidae) para identificar hábitats invernales y costos energéticos asociados con el comportamiento de grupo. Primero, comparamos densidades, la proporción sexo-edad y el uso de sustratos de forrajeo dentro y entre especies en función de las condiciones ambientales en 2 paisajes con diferente proporción de cobertura forestal. Posteriormente, comparamos el área de distribución y el esfuerzo de forrajeo para identificar costos energéticos asociados al comportamiento de grupo. Las densidades y proporción de sexo-edad fueron diferentes para las 2 especies entre sitios de alta y baja calidad. El uso mayoritario de substratos interiores como ramas y troncos y estratos bajos en sotobosque fue observado en del chipe de Townsend, mientras que el chipe de cachetes amarillos utilizo con más frecuencia substratos externos como hojas, ramas inmaduras y el dosel. El ámbito hogareño del chipe de Townsend mostró superposición temporal con el área utilizada por parvadas sugiriendo potenciales beneficios adicionales al retener un territorio. La distancia de viaje, área de uso diario y tiempo de forrajeo fue mayor en sitios de menor calidad para ambas especies siendo el tamaño de grupo un factor determinante.
Palabras clave: Competencia; Chipe cachetes amarillos; Parvadas mixtas forrajeras; Neotrópico
Introduction
Territoriality during the breeding period is an adaptive strategy of space use that most warbler species (Parulidae) use to secure habitats where reproductive success and survival rates are higher (Brown, 1969; Holmes et al., 1989; Sih, 1980; Stephen & Krebs, 1986). Conversely, during the nonbreeding period, alternative strategies of space use such as floating and group behavior (e.g., coordinated foraging among individuals in a species or between 2 or more different species), among others, are also selected to maximize fitness and reduce carry-over effects (i.e., lower breeding success due to delayed arrival or poor body condition) in the following reproductive season (Brown & Sherry, 2008; Greenberg & Salewski, 2005; Lima & Dill, 1990; Norris, 2005; Winker, 1998). While trait-biased habitat segregation still occurs in the wintering grounds with adult males or females occupying higher-quality habitats (Marra, 2000; Marra et al., 2015; Stutchbury, 1994), group behavior is more common as the dry season progresses in the wintering grounds (Brown & Sherry, 2008; Greenberg & Salewski, 2005; Powell, 1985). Transfer of social information on resource location and predation risk drives aggregation of monospecific or mixed-species bird flocks (Gil et al., 2017; Goodale et al., 2010; Valone, 2007), where additional fitness benefits, such as higher foraging efficiency and reduced predation rates, may be accrued by flock attendants, including former territory owners that join flocks (Dolby & Grubb, 1998; Krause & Ruxton, 2002; Sridhar et al., 2009; Terborgh, 1990). Group behavior is widespread in the wintering neotropical grounds (King & Rappole, 2000), but the net per-capita effect of flocking on fitness often varies among attendants (e.g., between dominant or more experienced and competitive adult individuals and subordinate or first year juvenile individuals), an outcome that is mediated by group size and environmental conditions (Goodale et al., 2020; Mokross et al., 2018; Pollard & Blumstein, 2008; Smith, 1976).
In the Mesoamerican Region, pine-oak forest environments are facing accelerated rates of deforestation due to chronic extraction of wood, and land clearing for agriculture, livestock grazing and housing development (Ramírez-Marcial et al., 2001). Habitat loss and degradation have reduced the availability of wintering areas used by migratory birds, forcing individuals into lower quality habitat as a function of higher occupancy in habitats where seasonal survival is higher (Morris, 1988; Steven et al., 2020). Weather seasonality in addition to land use change also limit forest insect productivity which in turn promote shifts in strategies of space use and potential overlap between otherwise segregated species or individuals (Janzen, 1980; Kent et al., 2022). To maximize fitness, warbler species may opt for the least costly strategy that provides the most benefits (e.g., defending a territory), but this may depend on the quantity and quality of available habitats as well as additional opportunities provided by others such as mixed-species flocks (Powell, 1985; Smith et al., 2010; Sridhar et al., 2009). When territory quality changes and resources are limited, aggregation in flocks often increase as a function of habitat degradation and resource location when prey density is sparse (Jullien & Clobert, 2000; Munn & Terborgh, 1979). However, a trade-off between finding profitable food sites and increased foraging costs, such as competition and higher foraging effort, may be possible when groups are larger and forage in “risky” areas without protective cover or along forest edges (Hutto, 1988; McNamara & Houston, 1992; Mokross et al., 2018; Pomara et al., 2007; Telleria et al., 2001). Foraging in open areas is known to increase perceived predation risk in forest birds and exerts non-lethal effects on flocks by inducing changes to group size (Laundré et al., 2010). On the other hand, a larger flock group may spend more time foraging as they are less vigilant and would establish social hierarchies as food resources may be quickly depleted (Sridhar et al., 2009).
While differential strategies of space use and patterns of habitat selection in wintering populations results in differential survival rates for most territorial warbler species (Greenberg & Salewski, 2005; Steven et al., 2020), limited data on social structure of species during flocking may also constrain development of strategies to protect them in their non-breeding grounds. Despite observed changes in species composition, sex-age structure, and social dynamics of flocks under changing habitat conditions (Gentry et al., 2022; Gil et al., 2017), it remains less known about how closely relative species that join flocks differ in their habitat use as a function of site quality (e.g., undisturbed vs disturbed) and how this also reflects the variation in foraging effort as an indirect proxy for fitness.
Our main objective was to assess patterns of habitat use at the micro (foraging substrate/layers) and meso (forest stand) levels within and between 2 congeneric migratory warbler species that coexist in flocks during the non-breeding period. Our secondary objective was to assess foraging effort as a function of group behavior measured through proxies on energy expenditure to determine whether differences in these responses also vary within and between the 2 species. Our overall goal was to provide evidence on changes in spatial variation in habitat use and foraging effort as a function of site quality. Assessing these responses is informative to identify wintering areas and to identify costs associated with changing environments, aiming to tailor specific interventions in forest stand structure and composition in landscapes where populations of migrant species including threatened species coexist with the local resident avifauna.
To achieve the objectives, we used field observations on foraging substrate and stratums use by both the focal species and flocks. In addition, we used telemetry to estimate home range in some individuals to confirm if their territories overlap with flock occurrence as a potential strategy to maximize fitness. We then determined the potential effects on energy expenditure proxies that might result for flocking individuals, specifically foraging effort measured by daily travel distance, size of daily use areas, and cumulative searching times during foraging. Outcomes were explored across 2 study sites in Chiapas, México, one in primary forest habitat, “Moxvquil”, with lower anthropogenic disturbance and one suboptimal or secondary forest, “Encuentro”, with higher anthropogenic disturbance (Komar et al., 2011). We predicted differential use of substrates within and between species to limit negative interactions with adult males and females preferentially occurring in areas of flock occurrence, whereas juveniles will occur in the most disturbed habitats and at a lower proportion in flocks. Lastly, we predicted that lower density of trees, canopy cover and smaller height and diameter at breast height (D. B. H.) of trees at Encuentro should force longer searching times and travel distances, as well as larger daily use areas. Opposite responses were expected at the low-disturbance site, Moxviquil.
Materials and methods
In this research, we used as a study system 2 Neotropical-Nearctic migratory warbler species (family Parulidae), the Golden-cheeked warbler (Setophaga chrysoparia), hereafter GCWA, and the Townsend’s warbler (Setophaga townsendi), hereafter TOWA, that co-occur in flocks during their non-breeding period. The former species breeds exclusively in mature oak-juniper woodlands of central Texas (Wahl et al., 1990), whereas TOWA breeds in coniferous forests of northwestern USA (Wright et al., 1998). Both species overwinter in the Mesoamerican pine-oak forest ecoregion that extends from southern México to northwestern Nicaragua, a common nonbreeding area for warblers and a region where a steady decline in abundance of most migratory species is partially explained by habitat loss and degradation (Birdlife International, 2020). GCWA shows latitudinal sexual segregation and occurs at low densities across its nonbreeding grounds (Rappole et al., 2000), but underestimation of densities needs further confirmation as difficulty in distinguishing plumage characteristics may confound current estimates (Komar et al., 2011). TOWA co-occurs at similar elevations (1,200-2,900 m) and environments (pine-oak forests) although its winter range of distribution is larger than its close relative, GCWA, during their non-breeding periods (Wright et al., 1998). Both species join other migrants and residents in mixed-species flocks as a foraging strategy to deal with the variation in resource supply and predation risk throughout their wintering areas (Vidal et al., 1994). GCWA is considered an oak tree specialist and is listed as endangered in both the IUCN Red List Category and the NOM-059-SEMARNAT 2010 (Birdlife International, 2020; King et al., 2012; Semarnat, 2010). Conversely, TOWA shows little specialization in foraging, occurring in both primary and secondary habitats, but also forage in oak during the nonbreeding season (Greenberg et al., 2001). Being congeneric, the 2 warbler species share ecological and morphological similarities, and may compete for resources such as prey located on same foraging substrates (Losin et al., 2016; Newell et al., 2014). To date, no study has assessed microhabitat preferences and foraging effort during the non-breeding period, excepting for their occurrence in flocks (King & Rappole, 2000; Komar et al., 2011; Vidal et al., 1994).
Surveys were carried out in 2 study sites within the municipality of San Cristóbal de las Casas, in the state of Chiapas, México. Site 1 (Moxviquil) is a nature reserve (privately owned by Pronatura Sur) with an area of 101 ha and average altitude of 2,200 m, where the dominant vegetation types are second-growth oak forest and pine-oak associations. Closed canopy forest dominates and the density of trees, especially oaks, is relatively high compared to other landscapes where the Golden-cheeked warbler overwinters (Vidal et al., 1994). Moxviquil is assumed to be a primary habitat (higher quality) for the Golden-cheeked warbler where survival may be higher compared to lower-quality habitat where disturbance have simplified stand structure and composition along with open areas (Komar et al., 2011). Site 2 (Encuentro) comprises remnant patches of pine-oak forest within land used largely for grazing with semirural scattered housing settlements. The forest in Encuentro has been used mainly for wood extraction, an activity that creates open areas and forest stands with noncontinuous canopy cover and a lower density of trees than at Moxviquil. Encuentro is considered a secondary habitat (lower quality) with lower density for the Golden-cheeked warbler (Komar et al., 2011).
Vegetation data were obtained in each study site: Moxviquil (low disturbance) and Encuentro (high disturbance). We established 10 plots (0.1 ha in size) in the 2 study sites (Moxviquil, n = 5, and Encuentro, n = 5) where we measured the following variables per plot: number of oak tree species larger than 5 cm at diameter breast height, D. B. H. (diameter at breast height) of trees measured, oak tree height in meters using a Suunto clinometer, tree canopy cover percentage using a spherical crown densiometer, and total abundance of epiphytes per plot by counting the number of individuals per tree. Epiphytes were included as a known reservoir of insect prey for many bird species (Nadkarni & Matelson, 1989). All variables were measured using standard techniques in forest mensuration (Kershaw et al., 2017). To compare vegetation characteristics between Moxquivil and Encuentro, we used analysis of variance (ANOVA) for number of trees, D. B. H., height per site and a Kruskal Wallis test for the variables with data not normally distributed (i.e., epiphytes and canopy cover).
Survey seasons ran from November to March when migrants were assumed to be occupying winter habitats and not migrating farther southward. Sampling was carried out in one season: Nov. 2018 – Feb. 2019, but additional information from 2 previous seasons (Nov. 2016 – Feb. 2017 and Nov. 2017 – Feb. 2018) was provided by a local organization (Pronatura Sur). This was aimed to increase sample size to obtain an average metric of flock group size and foraging effort proxies through seasons and per site. Information provided by Pronatura Sur was obtained using the same field methods as in this study.
Based on confirmation of previous occurrence of the 2 study warbler species in 2 pine-oak forest sites, we selected 5 plot locations per site (Moxviquil, n = 5 and Encuentro n = 5) where flocks have been observed continuously through years (Pronatura Sur, pers comments). In each study site and plot we started the surveys by locating the flock formation at dawn (06:00 local time) to match the peak activity of birds. Once a flock (i.e., a group of individuals of at least 2 or 3 different species moving and foraging together) was located, we followed them until members abandoned the formation (usually 2 to 3 h on average). During the surveys we remained silent at a safer distance of 50 m to avoid interfering with their activity and counted and classified every GCWA and TOWA individual by sex and age (i.e., immature or first year, second year and adult) within a flock to obtain densities as well as flock group size. We were sure to count every individual even when flocks where foraging in the tree canopies whose heights were between 10 and 13 m on average. While flocks were moving across the landscape, flock speed was not an issue to follow them throughout terrain. To estimate densities of GCWA and TOWA per plot and year, we divided the number of individuals in a flock by its Minimum Convex Polygon (MCP) area in hectares. Normality assumptions on density data distributions were assessed with chi-square goodness-of-fit tests. Density comparisons by habitat were based on paired samples t-tests with an alpha value at 0.1 to acknowledge low sample size and improve the power of the tests.
Data on foraging traits for each individual were obtained through focal observations in periods of 60 s when warblers were spotted in flocks. To describe foraging by these individuals, we counted frequencies of observations on various substrate locations (outermost leaves, outer twigs 2-4 m from trunk, inner major branches 0-2 m from trunk, and trunk) and heights within the tree canopy (vertical layers: in shrubs or 0-1 m from the forest floor, in the understorey 1-2 m from the forest floor, at mid-level 2-8 m from the forest floor, and in the tree canopy), expressed as percent frequency of use for each substrate and vertical layer (Sodhi & Paszkowski, 1995). To compare the use of foraging substrates and vertical forest layers between species in both study sites and between sites we used general lineal models with family = binomial and a link function = logit to assess differences as a function of site and species.
We collected data on these metrics by using the same flocks we surveyed to determine flock characteristics (sex and age and flock group size). During the 2018-2019 survey season, we first located the foraging flocks and confirmed the occurrence of the 2 focal warbler species within them. We then followed each flock by visually tracking and recording their geographic positions every few minutes using the “tracklog” function of a GARMIN eTrex 22x GPS unit (precision ± 5 m, 40 to 60 location points per hour period where number of points varied as a function of flock aggregation time). The coordinate points dataset for each flock was used to obtain daily travel distance, calculated here as the sum of distances in meters between each geographic individual point in the data set. To estimate size of daily used areas, we took readings every 5 minutes with a different GPS unit to delineate a MCP per flock, with area expressed in hectares and estimated as the total area in size divided by the number of flocks in each site (Worton, 1989). Given that it was not possible to differentiate between feeding and moving time, we obtained the average cumulative searching time, estimated here as the total time spent by a flock while traveling until flock was disbanded. The 3 metrics: flock travel distance, daily used areas, and searching time, were considered as proxies for foraging time. Each metric was pooled across all surveyed flocks by season (3 seasons in total: 2016-2017; 2017-2018; 2018-2019) and study site to obtain mean and standard error that were compared between sites but not between years as there was no difference in each metric estimate across years. To compare travel distances, daily use areas and cumulative searching times between sites we used ANOVA. Then we used a generalized linear mixed model with family = gaussian, and a link function = identity with survey year (visit) as a random variable and group size as explanatory variable to assess if effort (cumulative searching time) varied as a function of these variables. We selected this latter variable as independent in the GLMM as the daily travel distance and daily used area were highly correlated (r = 0.88, N = 30, p < 0.001). All differences between vegetation variables and associations between variables used in ANOVA and the GLMM were considered significant at α ≤ 0.05. Analyses were performed with JMP® statistical software and R.
We obtained a banding permit from the Secretary of Natural Resources and the Environment (permit: 09/K5/–0127/01/20: SEMARNAT in Mexico) and approval from the Animal Care Committee at Lakehead University respecting the Canadian Council on Animal Care (CCAC) guidelines for Animal Use Protocol (permit: #11 2018, Romeo # 1466788). CCAC approval is a university requirement for field work involving research using live animals, even outside Canada. To capture birds in the Moxviquil site, we selected the same locations used to estimate group size, sex-age ratios and foraging substrates/layers use. However, search effort was limited only to this site due to logistic and safety constrains in the Encuentro site which was not sampled for this purpose. In the 5 locations/plots we placed mist nets to capture birds belonging to the 2 focal species. To increase capture probabilities, we used audio playbacks of their calls when they were likely in defended territories. Nets were placed during the month of February 2020, when the focal species were still occupying the site and we followed the individuals for 21 days due to battery transmitter duration. Sampled individuals were fitted with nano-transmitters (Model LB-2X, Holohil Systems Ltd.) using a harness made from elastic thread. Transmitter weight was 0.31 g, equivalent to less than 5% of the weight range in the 2 species (9-15 g), aimed to avoid movement constraints and unnecessary energetic demand. We tracked the individuals with a Telonics R-4 receiver and a hand-held antenna. Tracking began 24 h after capture and ran continuously until the battery life of the transmitter ended (7 weeks on average). With all tracked point locations, we then estimated minimum convex polygons for the 5 flocks and Kernel Utilization Distributions (KUD) in Moxviquil to generate 95% volume distribution contour for the home range and 50% volume distribution contour for the core foraging areas and potential territories in those individuals tagged with radio transmitters (Worton, 1989). This information was used to assess the spatial overlap between each MCP per flock and individual home ranges obtained from tagged individuals while foraging in flocks. Open JUMP GIS software (Steiniger & Hunter, 2012) was used to perform the home range analysis and estimation.
Results
Vegetation differed between sites, but not between plots at each site. Moxviquil forest plots were higher in density of mature oak trees, higher average tree height, denser canopy cover, and higher abundance of epiphytes than in Encuentro (Table 1).
| Variables | Moxviquil | Encuentro | Test statistic | p |
|---|---|---|---|---|
| Number of trees per site (1,000 m2) | 111 ± 1.3 | 47 ± 0.8 | F1,566 = 38.8 | < 0.001 |
| Diameter at breast height (cm) | 16.9 ± 0.2 | 18.3 ± 0.7 | F1,566 = 444.8 | < 0.001 |
| Average height of oak trees (m) | 13.8 ± 0.2 | 12.6 ± 0.7 | F1,566 = 4.34 | 0.037 |
| Mean abundance of epiphytes* | 2,780 ± 650 | 480 ± 70 | χ2 = 17.7 | < 0.001 |
| Canopy cover (%)* | 80.0 ± 0.2 | 63.6 ± 0.5 | F1,566 = 740.4 | < 0.001 |
Mean number of individuals per flock did not differ between plots at either site, but flocks were larger at Encuentro (35 ± 3 individuals) than at Moxviquil (27 ± 2 individuals; F1,28 = 5.64, p = 0.02). Group size varies from 13 to 40 individuals in Moxviquil and 20 to 45 in Encuentro. The total number of species in the mixed flocks was similar for Moxviquil (15 species) and Encuentro (16 species). The most abundant other species in flocks were the hermit warbler (Setophaga occidentalis), crescent-chested warbler (Oreothlypis superciliosa), hutton’s vireo (Vireo huttoni), and blue-headed vireo (Vireo solitarius), all forest dependant migratory species. While flocks were numerous at both sites, focal species relative abundances were lower, i.e., 1 or 2 GCWA individuals per flock and 3 to 5 TOWA individuals per flock in the more disturbed site (Encuentro) when compared with Moxviquil. When assessing densities of the focal species in this study, we only found 16 GCWA individuals at Moxviquil and 11 at Encuentro in 3 pairs of flocks (n = 5 per site × 2 sites = 10, total 30 flocks) across the 3 years. GCWA densities were not different between its preferred (0.71 ± 0.14 ha-1) and secondary habitats (0.65 ± 0.12 ha-1) at Moxviquil (t8 = 0.386, p = 0.71). TOWA densities were also not different between habitats at Moxviquil (3.1 ± 0.6 ha-1 and 3.2 ± 0.7; t8 = 0.11, p = 0.99). In Encuentro, densities of GCWA were higher in its preferred habitat (0.67 ± 0.1 ha-1) than in its secondary habitat (0.37 ± 0.7 ha-1; t8 = 3.28, p = 0.01). TOWA densities did not differ between habitats in Encuentro (2.5 ± 0.7 ha-1, 1.7 ± 0.4 ha-1; t8 = 1.24, p = 0.24).
Across 3 survey seasons at Moxviquil, 16 GCWA individuals were found. Nine were males, 7 were females. Eleven were adults (9 males and 2 females), and 5 (all females) were immatures. At Encuentro, 10 adult GCWA males alongside only 1 immature female was found. At Moxviquil, 53 TOWA individuals were found, 28 males and 25 females. Age was assessed in only 32 of these individuals because plumage similarities between juvenile male and females made assessment of the other 21 individuals difficult; 29 were adults and only 3 were immatures. Fourteen were adult males, and 15 were adult females. In the immature category, 2 were males, and 1 was a female. At Encuentro, 54 TOWA individuals were found, 37 males and 17 females. Forty-three of these were aged; 40 were adults, 1 an immature male and 2 immature females. In sum, for both focal species, age ratios differed (adults dominated) at Moxviquil, while both sex and age ratios differed (males and adults dominated) at Encuentro (Table 2).
| Variables | Moxviquil | Encuentro |
|---|---|---|
| GCWA sex ratio | 1.2 : 1 m (9): f (7) | 9 : 1 m (10): f (1) |
| GCWA age ratio | 2.2 : 1 a (11): imm (5) | 9 : 1 a (10): imm (1) |
| TOWA sex ratio | 1.1 : 1 m (28): f (25) | 9.6 : 1 m (37): f (17) |
| TOWA age ratio | 2.1 : 1 a (29): imm (3) | 4 : 1 a (40): imm (3) |
Sample sizes of individuals in Moxviquil were n = 16 GCWA and n = 16 TOWA, whereas in Encuentro, they were n = 11 GCWA and n = 11 TOWA. Foraging observations in the GCWA corresponded to oak tree species use (more than 90% in Quercus spp.) whereas TOWA used oak and pine tree species (70% and 30% in Pinus sp.). Proportional use of foraging substrates such as twigs and trunks differed between species, but not between sites (Fig. 1, B = 1.46, SE = 0.64, z = -3.6, p = 0.007; B = 2.9, SE = 1.09, z = 2.6, p = 0.008). However, use of substrates such as inner branches differed between species and between sites (B = 1.7, SE = 0.62, z = 2.7, p = 0.006; B = -1.26, SE = 0.64, z = -1.98, p = 0.04) being this effect most notably at Encuentro. When referring to vertical layers use, differences in use of understory and canopy were found between species in both sites (B = 3, SE = 0.71, z = 4.2, p < 0.0001; B = -3.04, SE = 0.95, z = -3.2, p = 0.001) where proportional use of the understory was higher in TOWA, while on the other had, proportional use of canopy was higher in the GCWA in both sites (Fig. 2).
Daily travel distance by flocks differed between sites (F1,28 = 5.7, p = 0.02), being lower at Moxviquil (1.02 ± 0.80 km) than at Encuentro (1.30 ± 0.12 km). Daily used areas were statistically different (F1,28 = 5.1, p = 0.03) between Encuentro (3.27 ± 2.94 ha) and Moxviquil (2.03 ± 0.16 ha), but 2 areas at Encuentro were much larger (6.9 and 12.3 ha) than the average size at this site. Cumulative searching time differed between sites (F1,28 = 6.6, p = 0.01) and where lower at Moxviquil (160 ± 8 min) than at Encuentro (194 ± 11 min; Table 3).
| Variables | Moxviquil | Encuentro | Test statistic | p |
|---|---|---|---|---|
| Daily travel distance (meters) | 1.02 ± 0.8 | 1.3 ± 0.1 | F1,28 = 5.7 | 0.02 |
| Daily used area (hectares) | 2.03 ± 0.1 | 3.27 ± 2.9 | F1,28 = 5.1 | 0.03 |
| Cumulative searching time (minutes) | 160 ± 8 | 194 ± 11 | F1,28 = 6.6 | 0.01 |
The GLMM showed that cumulative searching time only increased as a function of flock group size (B = 3.29, SE = 0.44, t = 7.47, p < 0.001), and accounting for this effect in the GLMM resulted in significant differences in cumulative searching times across the 2 study sites although no effect of survey year and vegetation variables was related to this effort variable searching time (Table 4).
| Term | Coefficient | Std. error | t | p |
|---|---|---|---|---|
| Intercept | 77.26 | 37.92 | 2.03 | 0.05 |
| Site | – 9.32 | 32.18 | – 0.29 | 0.77 |
| Flock group size | 2.40 | 0.69 | 3.48 | < 0.01 |
| Flock size × site | 0.53 | 0.95 | 0.56 | 0.58 |
| Canopy cover percentage | 0.14 | 0.50 | 0.28 | 0.77 |
| Number of oak trees | 0.03 | 0.27 | 0.12 | 0.90 |
| Average tree height (m) | 5.81 | 10.75 | 0.54 | 0.59 |
| Mean number of epiphytes | – 0.39 | 0.58 | – 0.68 | 0.50 |
Fifteen adults were captured (TOWA, n = 14; GCWA, n = 1); 10 of 14 TOWA were males and the GCWA was a female. Median daily use area for all tracked individuals was 4.5 ± 0.8 ha; median home range (95% KUD) and core area (50% KUD) were 4.6 ± 0.7 ha and 1.4 ± 0.3 ha, respectively. Male home ranges (95% KUD for 10 captures) were smaller (3.5 ha, 1.6-5.5 ha 95% CI) than those of females (4 captures; 6.8 ha, 4.1-9.6 ha; t = 2.40, p = 0.03). Spatial overlap between flocks and home ranges was observed in 4 out of 5 flocks whereas temporal overlap throughout a larger survey window needs to be completed. The GCWA individual and 11 TOWA individuals out of 14 (80%) overlapped their home range either with 1 or 2 flocks, and the mean area of overlap was 0.76 ± 0.20 ha. Each home range overlapped a mean of 4.3 other home ranges, and individuals overlapped their home range more often in 2 of the 4 flocks observed (Fig. 3).
Discussion
Both focal warbler species were found together in mixed flocks at both sites. Flocks were larger at Encuentro, as predicted for a more disturbed landscape. For both species, adults were more commonly found in flocks, with males and females in equal proportions at Moxviquil whereas males were mostly found at Encuentro flocks, suggesting that intraspecific competition between males and females or with juveniles within flocks may be stronger in the lower-quality habitat (Encuentro) as seen for other groups where social hierarchies change once when in flocks (Darrah & Smith, 2014; Krebs, 1973; Martínez & Robinson, 2016). All tagged individuals with transmitters also occurred regularly in the Moxviquil flocks, despite theory predicting advantages of securing territories during unfavourable seasons (Marra et al., 2015), here, the relatively dry winter season. However, while this may hold for some species, plasticity in foraging behavior and strategies of space use is widespread in most migratory species (Greenberg & Salewski, 2005). Due to sample size and evidence from 1 location site, this assumption of territoriality needs further research to confirm if territorial individuals defend areas where flocks occur and throughout longer periods. While we predicted higher densities of juveniles in the Encuentro site, GCWA females and juveniles were found in similar proportions at both sites, although GCWA males were more abundant than females and juveniles in flocks. Some researchers have suggested latitudinal segregation in GCWA, where males occupy the northernmost nonbreeding grounds and females occur as far south as Nicaragua (Vidal et al., 1994). However, male abundance is also higher within the southernmost sites suggesting potential competition between males and juveniles for flock positions as it occurs in most tropical flocks, where hierarchies are established and whose impacts on foraging rates often determine foraging success on attendants (Goodale et al., 2020; Sridhar et al., 2011). In this study, only 1 adult GCWA female was captured and continuously observed in the same territory at Moxviquil during the duration of transmitter battery (21 days), whereas higher ratios of GCWA male adults in flocks of both sites also confirmed an increased abundance of adults when compared with females and juveniles. While it is possible that male adults are more common than females and juveniles when in flocks, other warbler species have shown equal proportions between males and females in flocks as observed in TOWA (King & Rappole, 2000). Further research is needed to confirm if same adult-juvenile proportions hold across latitudes or altitudes and if similar plumages of juveniles GCWA and TOWA may confound identification by observers. This is relevant because differential survival rates between adults and juveniles in a population through restricted access to juveniles by adults may determine population size and potential carryover effects (Marra et al., 2005; Norris, 2005).
When observing foraging responses at the substrate level, TOWA did not show any preference for a specific substrate and vertical layer. However, its foraging preferences over tree mid-level layers, may suggest an overlap and potential competition for microhabitats and insect resources with GCWA, influencing thereby shifts in habitat use of the latter species, as shown by Greenberg et al. (2001) for the black-throated green warbler (Setophaga virens). However, confirming the observed behavior of GCWA when TOWA is absent may show that competition is not a factor that determines foraging behavioral traits in the GCWA. While additional species in the flock may overlap in habitat use with GCWA, a test on trait similarity could also show if interspecific interactions had a displacement effect (e.g., through interference competition) on GCWA behavior.
As predicted, GCWA had “risky” foraging traits associated with a higher number of outermost leaves and twigs maneuvers and proportional use of canopy by GCWA at Encuentro, demonstrated through differential use of these substrates when compared with Moxviquil. Whether these behaviors potentially relate to resource partitioning, a specialization on specific insect prey are an evolutionary (morphological and behavioral) adaptation in response to past competition remains to be tested (Kent et al., 2022; Morris, 1988). On the other hand, GCWA used inner branches, trunks and the understory (forest floor/near shrubs) at a lower proportion at the disturbed site, suggesting that at Encuentro, TOWA may better defend or occupy these microsites that pose less predation risk in a disturbed forest landscape (Whittingham & Evans, 2004). While differential microhabitat use in the GCWA can be also attributed to competition by other species in the flock, TOWA may be more prone to compete with GCWA due to its ecological and morphological similarities. Despite observed niche differentiation between GCWA and TOWA, dominance hierarchies within flocks are known to reduce prey harvest rates if individuals increase vigilance and reduce foraging rates when they are displaced to the group periphery (Hamilton, 1971; Krause, 1994; Smith, 1976; Suhonen et al., 1993). Moreover, habitat partitioning and differences in foraging traits between species, has been described as an outcome of competition in areas with limited resources and higher predation risk (Greenberg et al., 2001; Martínez & Robinson, 2016). Given that resource partitioning also allows for coexistence of closely related species in resource limited environments (MacArthur, 1958), it may be necessary to measure prey abundance and differences in survival within and between species in habitats of varying quality as an extension of this study.
Group size potentially influenced effort and, thereby, the amount of time GCWA and TOWA spent foraging in flocks, especially in the more disturbed landscape. For forest-dwelling birds, a reduction in forest patch area and canopy cover, along with an increase in the distance between forest patches, may exert non-lethal effects on flocks by inducing changes to group size as seen in the more disturbed site, Encuentro (Gil et al., 2017; Laundre et al., 2010; Maldonado-Coelho & Marini, 2004; McNamara & Houston, 1992; Pollard & Blumstein, 2008). While larger groups are an immediate response to perceived predation risk in areas where forest structure and composition is less complex than in primary forests (Laundré et al., 2010; Telleria et al., 2001), a bigger group, on the other hand, may affect prey harvest rates if interspecific competition is higher between phylogenetically related flocking species whose niches overlap (Kent et al., 2022). However, to observe this response an assessment of agonistic bouts (e.g., chasing) and substrate use/dietary overlap needs to be recorded to confirm interference and exploitative competition. Nevertheless, an increase in flock size should still provide information on prey location as a function of disturbance at a site like Encuentro, where more competitors may still deplete foraging areas more rapidly (Gil et al., 2017; Hutto, 1988). The opposite may be expected in non-territorial systems or in systems with less competition such as in Moxviquil, where flocks were smaller and sex-age ratios were more even. Foraging effort also seemed to increase with flock size, as daily use areas, travel distances and searching time were larger or longer in Encuentro. However, a merely increase in group size may not be an indicative of higher foraging effort as vegetation characteristics at the landscape level such as patch area size and distance between forest fragments may have a bigger impact on foraging behavior (Mokross, 2018; Telleria, 2001). The increase in foraging effort proxies at Encuentro is proposed here as an indirect effect of group size which varied between sites due to the potential effect of changes in vegetation structure, where reduced density of trees and lower canopy cover forces groups to allocate more time to searching for food over longer distances (Laundré et al., 2010; Mokross et al., 2018). While larger flocks provide early detection and protection against predators through mobbing calls and risk dilution (Elgar, 1989; Valone, 2007), the less abundant species in a mixed flock may be also more conspicuous to predators through increased apparent competition with the more abundant species (Holt & Lawton, 1993; Jullien & Clobert, 2000). Landscape and habitat conditions likely influenced the variation in foraging effort and potential energy expenditure in GCWA, but it seems that adults, found more often in flocks within the disturbed site Encuentro, are trading off foraging efficiency and higher threat of predation for larger group sizes (Gil et al., 2017; Krause & Ruxton, 2002). This response calls to test for additional effects on foraging effort on flocking individuals because dominants are assumed to secure the higher quality habitats with reduced foraging costs (Marra et al., 2015).
Most captured TOWA individuals occupied home ranges coincident with the territory of one or more flocks. Home ranges that overlapped flock territories were smaller than those that did not, suggesting that joining flocks does not require a larger home range. However, given the smaller sample size on captured individuals for GCWA, this affirmation should be taken with precaution. The smaller ranges, occupied by mostly adult males, presumably more competitive individuals, may be the outcome of a recognized benefit of maintaining a fixed territory when a flock shows site fidelity or occurs in the same areas across different seasons as shown previously (Darrah & Smith, 2014). Tropical bird flocks have been also shown to be stable in composition and size, occupying and defending the same foraging areas through several seasons (Martínez & Robinson, 2016; Thomson et al., 2003). A further interpretation is that juvenile individuals do not have a flock near their home range, or do not hold a home range, as relatively few females and no immatures were captured during the study. Thus, a dual strategy, in which territory-holders, also observed outside flocks may forage locally and alone or where they may also regularly opt for group foraging, may be the case for these warblers where variation in food supply occurs both spatially and temporally, especially where environmental changes also occur rapidly (Greenberg & Salewski, 2005). However, increasing the survey window (through years) and sample size in future study of the 2 warblers will allow to test for the idea that not only do some overwintering warblers adhere to territories, illustrating despotic behaviour (Marra et al., 2015; Stutchbury, 1994), but they also position their individual territories along with flock territories to limit the occupation of both by subordinates or less competitive individuals. Assessing this information through additional research is relevant because differential strategies of space use may translate into differential survival and contribution to population size that will carry over the next breeding season for most Neotropical-Nearctic migratory bird species.
Conservation implications. The association of the 2 focal species with different foraging substrates and vertical layers in this study is important to design forest management strategies that prioritize mature trees with higher structural and horizontal complexity and plant species composition at the forest stand level, where species find suitable substrates and stratums to forage. A strong association between GCWA and oak tree species also suggest this native species should be prioritized during reforestation activities to increase forest cover in the disturbed site Encuentro and to increase tree abundance in Moxviquil, where wood extraction is still present. Given that searching efforts were larger in Encuentro, potentially as a function of larger search areas, it is recommendable to increase connectivity and decrease open area size where flocks may reduce their travelled distances and daily home ranges. Managing habitat for additional species that compose flocks, mostly migrant species, as seen in the Moxviquil site through continuous protection, will also promote habitat occupancy for the GCWA, which seems to be an obligated flock member as all observations of individuals, but one, were made in flocks.
Acknowledgments
To Secihti México (former Conacyt) who provided financial support through a doctoral scholarship awarded to first author. To Pronatura Sur and “Alianza para la conservación de los bosques de pino-encino de Mesoamérica” staff including Eric Hernández Molina and Claudia Macías Caballero who allowed field work at the Moxviquil Reserve (SCLC, Chiapas) and shared personnel and valuable field data to complement this study.
| No. | Family | Species | English name |
|---|---|---|---|
| Moxviquil | |||
| 1 | Parulidae | Setophaga townsendi | Townsend’s warbler |
| 2 | Parulidae | Oreothlypis superciliosa | Crescent-chested warbler |
| 3 | Vireonidae | Vireo huttoni | Hutton’s vireo |
| 4 | Aegithalidae | Psaltriparus minimus | Bushtit |
| 5 | Vireonidae | Vireo solitarius | Solitary vireo |
| 6 | Parulidae | Peucedramus taeniatus | Olive warbler |
| 7 | Parulidae | Setophaga occidentalis | Hermit warbler |
| 8 | Parulidae | Mniotilta varia | Black and white warbler |
| 9 | Tyrannidae | Contopus pertinax | Greater pewee |
| 10 | Parulidae | Setophaga chrysoparia | Golden-cheeked warbler |
| 11 | Parulidae | Cardellina Pusilla | Wilson’s warbler |
| 12 | Furnariidae | Lepidocolaptes affinis | Spot-crowned woodcreeper |
| 13 | Parulidae | Cardellina rubrifrons | Red-faced warbler |
| 14 | Parulidae | Myioborus miniatus | Slate-throated redstart |
| 15 | Certhidae | Certhia americana | American woodcreeper |
| Encuentro | |||
| 1 | Parulidae | Setophaga occidentalis | Hermit warbler |
| 2 | Parulidae | Setophaga townsendi | Townsend’s warbler |
| 3 | Parulidae | Oreothlypis superciliosa | Crescent-chested Warbler |
| 4 | Vireonidae | Vireo huttoni | Hutton’s vireo |
| 5 | Vireonidae | Vireo solitarius | Solitary vireo |
| 6 | Certhidae | Certhia americana | American woodcreeper |
| 7 | Parulidae | Peucedramus taeniatus | Olive warbler |
| 8 | Furnariidae | Lepidocolaptes affinis | Spot-crowned woodcreeper |
| 9 | Parulidae | Cardellina Pusilla | Wilson’s warbler |
| 10 | Parulidae | Mniotilta varia | Black and white warbler |
| 11 | Tyrannidae | Contopus pertinax | Greater pewee |
| 12 | Parulidae | Setophaga chrysoparia | Golden-cheeked warbler |
| 13 | Tyrannidae | Mitrephanes phaeocercus | Tufted flaycatcher |
| 14 | Parulidae | Cardellina rubrifrons | Red-faced warbler |
| 15 | Parulidae | Setophaga virens | Black-throated green warbler |
| 16 | Parulidae | Myioborus miniatus | Slate-throated redstart |
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Modelo de nicho ecológico de Cedrelinga cateniformis en la Amazonía del Ecuador
Ecological niche modeling of Cedrelinga cateniformis in the Ecuadorian Amazon
Daniel Adrian Vistin-Guamantaqui a, *, Carmen Elena Mantilla-Cabrera b, Paul Marcelo Tacle-Humanante c y Sulaya Betsabe Bayancela-Delgado c
a Escuela Superior Politécnica de Chimborazo, Facultad de Recursos Naturales, Av. Pedro Vicente Maldonado s/n y Av. 11 de Noviembre, Riobamba, Ecuador
b Escuela Superior Politécnica de Chimborazo, Facultad de Recursos Naturales, Agronomía, Grupo de Investigación Seguridad Informática y Telemática, Av. Pedro Vicente Maldonado s/n y Av. 11 de Noviembre, Riobamba, Ecuador
c Escuela Superior Politécnica de Chimborazo, Facultad de Recursos Naturales, Grupo de Investigación de Manejo y Aprovechamiento de los Recursos Renovables, Av. Pedro Vicente Maldonado s/n y Av. 11 de Noviembre, Riobamba, Ecuador
*Autor para correspondencia: adrianvistin@yahoo.com (D.A. Vistin-Guamantaqui)
Recibido: 12 noviembre 2025; aceptado: 10 abril 2026
Resumen
La Amazonía ecuatoriana es uno de los ecosistemas más biodiversos del planeta, enfrenta presiones derivadas de la deforestación, expansión agrícola, minería y fragmentación del paisaje, que amenazan su resiliencia ecológica. Cedrelinga cateniformis (Ducke) Ducke destaca por su rápido crecimiento, adaptabilidad a suelos ácidos y capacidad de mejorar la fertilidad, constituyendo una alternativa para la restauración mediante sistemas agroforestales. El objetivo de este estudio fue modelar el nicho ecológico con el algoritmo MaxEnt en las 6 provincias amazónicas del Ecuador para identificar áreas de distribución potencial y apoyar la planificación de estrategias de conservación y producción sostenible. Se utilizaron 19 variables bioclimáticas tomadas de WorldClim más la elevación, validando 136 registros de presencia. Tras eliminar colinealidades (r > 0.70), se seleccionaron 10 variables; la validación (75/25%) se realizó con AUC y TSS. El modelo alcanzó un AUC de 0.992. La variable más influyente fue la precipitación del mes más seco (BIO14; 71.9%), identificándose 13,288 km² de idoneidad óptima (11.07% del territorio), principalmente en Sucumbíos y Orellana.
Palabras clave: Distribución potencial; Variables bioclimáticas; Idoneidad ambiental; Sistemas agroforestales
Abstract
The Ecuadorian Amazon, one of the most biodiverse ecosystems on the planet, faces increasing pressures from deforestation, agricultural expansion, mining, and landscape fragmentation, threatening its ecological resilience. Cedrelinga cateniformis (Ducke) Ducke is notable for its rapid growth, adaptability to acidic soils, and capacity to improve soil fertility, making it a promising species for ecological restoration through agroforestry systems. This study aimed to model its ecological niche using the MaxEnt algorithm across the 6 Amazonian provinces of Ecuador to identify potential distribution areas and support conservation and sustainable production strategies. Nineteen bioclimatic variables from WorldClim and elevation data were analyzed using 136 validated presence records. After removing collinearity (r > 0.70), 10 variables were selected, and the model was validated (75/25%) using AUC and TSS. The model showed excellent performance (AUC = 0.992). The most influential variable was precipitation of the driest month (BIO14; 71.9%). Approximately 13,288 km² (11.07% of the territory) showed optimal suitability, mainly in Sucumbíos and Orellana, highlighting key areas for restoration and conservation planning in the Ecuadorian Amazon.
Keywords: Potential distribution; Bioclimatic variables; Environmental suitability; Agroforestry systems
Introducción
La Amazonia del Ecuador abarca aproximadamente 2% de la cuenca del Amazonas, ésta se extiende por 6 provincias en el Ecuador (Sucumbíos, Napo, Orellana, Pastaza, Morona Santiago y Zamora Chinchipe). Este espacio es reconocido como una de las regiones más biodiversas del planeta, donde se establecen vastos bosques húmedos tropicales y una extraordinaria variedad de flora y fauna (Arias-Gutiérrez et al., 2016). Estos territorios han sido ocupados por indígenas de múltiples nacionalidades como son los Waoranis, Kichwa y Shuar, quienes por décadas han mantenido una relación ancestral y de dependencia con este ecosistema, aprendiendo a convivir en equilibrio con la naturaleza (Buitrón-Cañadas y López-Sandoval, 2019). Este ecosistema biodiverso genera un sinnúmero de servicios ambientales, principalmente el de proveer alimentos, medicinas, agua dulce, materiales de construcción para las viviendas, regulación climática y hábitats vitales para que se desarrollen múltiples organismos, es fuente de servicios culturales ligados a la cosmovisión y prácticas tradicionales ancestrales (Álvarez et al., 2024).
Alarcón-Aguirre et al. (2021) señalan que el bosque húmedo tropical del Ecuador es reconocido como un hotspot de biodiversidad a nivel mundial; sin embargo, actualmente enfrenta un alarmante proceso de deterioro debido a múltiples presiones de origen antrópico que comprometen su funcionalidad ecológica. Entre estas, la deforestación constituye el principal factor de impacto, impulsada principalmente por la apertura de áreas destinadas a la ganadería y al establecimiento de cultivos comerciales como palma africana (Elaeis guineensis), arroz (Oryza sativa), malanga (Colocasia esculenta) y maíz (Zea mays). Entre 2001 y 2020 se reportó la pérdida de más de 623,000 ha, concentradas principalmente en las provincias de Morona Santiago y Sucumbíos. Por otro lado, la actividad minera se ha consolidado como un motor adicional de degradación de este ecosistema, que provoca la destrucción de extensas áreas y afecta severamente la calidad del suelo y de los recursos hídricos, especialmente por la contaminación asociada al uso de mercurio y cianuro (Anto-Rubio, 2020).
De la misma manera, la fragmentación es un fenómeno que ocasiona una reducción de la viabilidad genética de innumerables especies, lo que ocasiona un impacto irreversible en los recursos genéticos, así como alteraciones en los ciclos ecológicos (Vargas-Chaves, 2024). Este conjunto de presiones se agrava debido a la falta de planes, programas y proyectos que viabilicen su restauración y resiliencia ecológica con la urgencia que requiere esta crisis, si bien en el Ecuador existe la iniciativa como es el Plan Nacional de Restauración del Paisaje 2021-2030 (Programa Nacional de Reforestación, 2019) y proyectos piloto que exploran un conjunto de métodos y técnicas en el ámbito de las micorrizas y su utilidad en la restauración (Arias-Gutiérrez et al., 2016), la escala y la velocidad de implementación son a menudo superadas por el avance de las actividades extractivas, entonces nace la necesidad de fortalecer la gobernanza a nivel local regional y nacional acompañada de la implementación de políticas públicas que se comprometan en desarrollar diferentes mecanismos factibles para la restauración ecológica activa y pasiva en sus territorios, principalmente en las zonas más afectadas (Conrado-Da Cruz et al., 2021).
Las comunidades oriundas de estos sitios, durante muchos años, practican el uso tradicional de la tierra conocido como chakra. En la provincia de Napo, esta práctica es promovida principalmente por las comunidades indígenas Kichwa y Kijus y ha alcanzado un reconocimiento a nivel mundial debido a su importancia cultural, productiva y ambiental (Álvarez et al., 2024). Tal es su relevancia, que la Organización de las Naciones Unidas para la Alimentación y la Agricultura (FAO) la ha reconocido como Patrimonio Agrícola Mundial. Este sistema productivo cubre una extensión aproximada de 162,082 ha y se caracteriza por la combinación de múltiples especies, principalmente cultivos de ciclo corto, como yuca (Manihot esculenta), plátano (Musa spp.), chonta (Bactris gasipaes), cacao (Theobroma cacao) y café (Coffea arabica), así como otras especies frutales y medicinales, lo que contribuye a la conservación de la biodiversidad y a la sostenibilidad del paisaje amazónico (Heredia-R et al., 2020).
Si bien no existe una cifra exacta en el Ecuador sobre la implementación de estos sistemas, de manera general fuentes gubernamentales manifiestan que son relativamente muy escasas debido a la presión de los monocultivos y modelos agrícolas comerciales, baja rentabilidad monetaria, limitaciones de mercado y cadenas de comercialización, falta de políticas públicas consistentes y apoyos gubernamentales. También otros factores sociológicos como es la migración, el cambio generacional y pérdida de conocimientos influyen en la presión sobre las tierras y conflictos por el uso del suelo (Uribe-Taborda et al., 2020).
Hernández-Núñez et al. (2021) manifiestan que los sistemas agroforestales (SAF) representan un sistema de uso integrado del suelo, se considera que están profundamente arraigados y científicamente validados como una alternativa viable en contra de los monocultivos, que son los que en su mayoría impulsan la deforestación y la degradación de la superficie arable (De Santana et al., 2023). Estos sistemas se implementan en diversas tipologías donde destacan los sistemas agrosilvícolas, los cuales integran cultivos con árboles nativos y los sistemas silvopastoriles que son una asociación de pastos y especies arbóreas (Kay et al., 2019). Un ejemplo es con C. cateniformis (Ducke) (Fabaceae), especie de árbol que se caracteriza por su alta adaptabilidad a suelos ácidos, arcillosos y de baja fertilidad, lo que la convierte en una especie especialmente relevante para programas de rehabilitación forestal, recuperación de suelos degradados y la implementación de sistemas agroforestales (Scarcelli et al., 2025). Asimismo, su madera presenta una densidad media y buena durabilidad, características que le confieren múltiples usos, entre ellos la carpintería, la construcción y la fabricación de muebles de alto valor económico (Haag et al., 2020).
Dentro del aspecto ecológico, esta especie contribuye a la captura de carbono, de igual manera mejora la fertilidad del suelo especialmente cuando se la implementa en sistemas silvopastoriles y agroforestales, ya que incrementa la macrofauna y la actividad biológica del suelo (Díaz-Pablo et al., 2024). Por otro lado, C. cateniformis presenta una ventaja en aspectos de propagación, ya que sus semillas poseen buena viabilidad facilitando su uso en programas de restauración ecológica debido a que tiene un crecimiento rápido y alta resistencia a plagas como el barrenador de brotes (Hypsipyla grandella) ya que se ha identificado que interactúa biológicamente con bacterias nitrificantes (García-Quintana et al., 2023). Estudios recientes y de campo han destacado que la especie enfrenta amenazas significativas en el Ecuador debido a la deforestación, fragmentación y la expansión de la frontera agrícola, motivo por el cual en los últimos años sus poblaciones se han visto disminuidas de manera alarmante, con efectos negativos en la provisión de los innumerables beneficios ecológicos y socioeconómicos asociados a esta especie. Con base en esta importancia nace la iniciativa de proponer a C. cateniformis como un eslabón clave para la gestión de la biodiversidad y resiliencia ecológica en el ecosistema tropical del Ecuador (Vistin-Guamantaqui et al., 2025).
Los sistemas de información geográfica (SIG) se han consolidado como herramientas fundamentales para la gestión de la biodiversidad, especialmente en el modelamiento de nichos ecológicos de especies arbóreas (Pauletto et al., 2025). A través de la integración de variables ambientales, climáticas y edáficas, los SIG permiten predecir la distribución potencial de las especies bajo condiciones actuales y escenarios futuros de cambio climático. Esto facilita la identificación de áreas prioritarias para la conservación y restauración de los ecosistemas (Franklin, 2023).
Diversos estudios han demostrado que el cambio climático y la deforestación reducen significativamente el hábitat disponible para numerosas especies, lo que evidencia la necesidad urgente de implementar políticas de conservación y restauración in situ y ex situ. En este contexto, Vieira-Capucho et al. (2025) señalan que los modelos de nicho ecológico constituyen una herramienta clave, ya que permiten identificar áreas con alta idoneidad climática y anticipar los efectos del cambio climático. De esta manera, estos modelos se convierten en una estrategia de manejo adaptativo que facilita la planificación y ejecución de programas y proyectos de conservación, revegetación, forestación y sistemas agroforestales. Asimismo, Morais-Sousa et al. (2025) destacan que este tipo de modelamiento también favorece la conectividad entre hábitats y fortalece la resiliencia de los ecosistemas forestales, permitiendo priorizar acciones de manejo y conservación orientadas a reducir la vulnerabilidad de las especies. En conjunto, estos enfoques contribuyen a la preservación de la biodiversidad, la cual resulta fundamental para mantener los servicios ecosistémicos esenciales para la región y el bienestar de las comunidades locales (Calvas-Serrano et al., 2024).
El objetivo de esta investigación fue desarrollar un modelo de nicho ecológico para C. cateniformis en las 6 provincias amazónicas del Ecuador mediante la utilización del algoritmo de máxima entropía, con el fin de identificar áreas potenciales de distribución que permitan proponer sitios adecuados para la gestión de sistemas agroforestales sostenibles y contribuyan a la conservación de la especie y a incrementar la resiliencia ecológica de los ecosistemas amazónicos del Ecuador.
Materiales y métodos
La zona de estudio abarca 6 provincias que conforman la Amazonía ecuatoriana, la cual se extiende desde la Cordillera Oriental de los Andes, al este hasta la frontera con Perú y el sur de Colombia. Esta región cubre una superficie aproximada de 120,000 km² y constituye una de las zonas más biodiversas y ecológicamente relevantes del país (López et al., 2013). Abarca 6 provincias, de las cuales Pastaza ocupa 29,628.8 km², Morona Santiago 24,029.1 km², Orellana 21,675.4 km², Sucumbíos 18,146.5 km², Napo 12,542.5 km² y Zamora Chinchipe 10,565.8 km² (fig. 1) (Vistin-Guamantaqui et al., 2025). La región se divide en 2 zonas altitudinales: la alta Amazonía, con elevaciones superiores a 1,000 m snm, y la llanura Amazónica, con altitudes inferiores a este nivel. Presenta un clima tropical muy húmedo durante todo el año, influenciado por la intensa evapotranspiración de los extensos bosques amazónicos.
Las precipitaciones son constantes a lo largo del año, con un ligero incremento entre marzo y julio y una disminución en los meses de agosto y enero, variaciones asociadas al desplazamiento de la zona de convergencia intertropical (ZCIT) (León-Baque et al., 2021). En la región amazónica, la precipitación alcanza valores superiores a 4,500 mm anuales, lo que la convierte en una de las zonas más lluviosas del Ecuador (INAMHI, 2025). Esta elevada pluviosidad se distribuye de manera relativamente uniforme a lo largo del año, contribuyendo al mantenimiento de los ecosistemas húmedos y a la densa cobertura vegetal característica de la zona, en cuanto al régimen térmico, la temperatura promedio anual oscila entre 24 y 25 ºC.
Los registros de la presencia de C. cateniformis fueron obtenidos mediante una revisión exhaustiva y sistemática de múltiples fuentes de información, que incluyeron colecciones de herbarios digitales, así como bases de datos en línea provenientes de servicios globales de información sobre diversidad biológica: Global Biodiversity Information Facility (GBIF, 2025), Missouri Botanical Garden (Tropicos, 2025). Se utilizó la plataforma Google Earth como herramienta de apoyo geoespacial para la verificación y validación de los puntos de presencia de C. cateniformis. Mediante la inspección visual de las coordenadas y la comparación con coberturas de referencia, se determinó la precisión espacial de cada punto de presencia. Como resultado, se consolidó un total de 136 registros georreferenciados, distribuidos en las 6 provincias que conforman la región amazónica del Ecuador, empleando como sistema de referencia geodésico el World Geodetic System 1984 (WGS 84) (fig. 1).
El modelo de distribución potencial de C. cateniformis fue elaborado a partir de variables bioclimáticas con una resolución espacial de 1 km (equivalente a 0.008983° ó 32.34”), obtenidas de la base de datos WorldClim (Hijmans et al., 2025). Este conjunto incluyó 19 variables bioclimáticas (tabla 1), las cuales describen patrones espaciales y temporales de temperatura y precipitación en la región de estudio, factores determinantes en la distribución ecológica de la especie. Con los resultados del modelo, tomando en cuenta la permutación y el porcentaje de contribución, se seleccionaron las variables bioclimáticas (BIOs) para la validación del modelo: (BIO1) temperatura media anual, (BIO2) rango medio diurno, (BIO8) temperatura media del trimestre más lluvioso, (BIO9) temperatura media del trimestre más seco, (BIO12) precipitación anual, (BIO14) precipitación del mes más seco, (BIO15) estacionalidad de la precipitación, (BIO17) precipitación del trimestre más seco y (BIO19) precipitación del trimestre más frío. Adicionalmente, se incorporó la elevación, también obtenida de WorldClim, con el fin de representar de manera más precisa los gradientes topográficos asociados a la distribución de la especie.
| Variables | Descripción | Unidad |
|---|---|---|
| BIO 1 | Temperatura media anual | °C x 10 |
| BIO 2 | Rango de temperaturas diurnas | °C x 10 |
| BIO 3 | Isotermalidad (BIO2/BIO7) (* 100) | % |
| BIO 4 | Estacionalidad en la temperatura (desviación estándar * 100) | °C x 100 |
| BIO 5 | Temperatura máxima del mes más cálido | °C x 10 |
| BIO 6 | Temperatura mínima del mes más frío | °C x 10 |
| BIO 7 | Rango anual de temperatura (BIO5-BIO6) | °C x 10 |
| BIO 8 | Temperatura media del trimestre más lluvioso | °C x 10 |
| BIO 9 | Temperatura media del trimestre más seco | °C x 10 |
| BIO 10 | Temperatura media del trimestre más cálido | °C x 10 |
| BIO 11 | Temperatura media del trimestre más frío | °C x 10 |
| BIO 12 | Precipitación anual | mm |
| BIO 13 | Precipitación del mes más lluvioso | mm |
| BIO 14 | Precipitación del mes más seco | mm |
| BIO 15 | Estacionalidad en la precipitación (coeficiente de variación) | % |
| BIO 16 | Precipitación del trimestre más lluvioso | mm |
| BIO 17 | Precipitación del trimestre más seco | mm |
| BIO 18 | Precipitación del trimestre más cálido | mm |
| BIO 19 | Precipitación del trimestre más frío | mm |
| ELEV | Elevación | m snm |
El desempeño del modelo de nicho ecológico fue evaluado mediante un conjunto de métricas ampliamente utilizadas en estudios de modelación de la distribución de especies. En primer lugar, se empleó el área bajo la curva (AUC, por sus siglas en inglés), la cual permite cuantificar la capacidad del modelo para discriminar entre áreas con condiciones ambientales favorables y no favorables sin depender de un umbral específico. De manera complementaria, se realizó una validación cuantitativa basada en la construcción de una matriz de confusión, utilizando 136 registros de presencia de la especie, los cuales fueron contrastados con las predicciones generadas por el modelo en la Amazonía ecuatoriana (Beeman et al., 2021).
La salida continua del modelo, expresada como un gradiente de idoneidad ambiental, fue binarizada en categorías de presencia y ausencia mediante la aplicación de un umbral de corte seleccionado en función del equilibrio entre sensibilidad y especificidad. A partir de la matriz de confusión resultante, se calcularon la sensibilidad —que expresa la capacidad del modelo para identificar correctamente las presencias reales— y la especificidad, que refleja su habilidad para discriminar las ausencias. Con base en estos parámetros, se estimaron la estadística de habilidad verdadera (TSS) y el índice de Kappa, métricas que permiten evaluar la concordancia entre las observaciones y las predicciones, descontando el efecto del azar. En conjunto, estas métricas posibilitaron una evaluación integral de la capacidad predictiva y la fiabilidad espacial del modelo de nicho ecológico desarrollado (Cruz-Román et al., 2023).
Se aplicó un análisis de componentes principales (PCA) utilizando las variables bioclimáticas (BIOs) con el propósito de identificar aquellas que más contribuyen a la presencia de C. cateniformis en la región amazónica del Ecuador. Este método multivariante permitió reducir la dimensionalidad de las variables y sintetizar la variabilidad ambiental en ejes ortogonales que representan los principales gradientes climáticos del área de estudio. Para la construcción del espacio ambiental se consideraron los 2 primeros componentes principales, los cuales explicaron la mayor proporción de la varianza total de los datos. Posteriormente, este espacio fue dividido en una cuadrícula de 100 celdas, con el fin de facilitar el análisis comparativo entre las condiciones ambientales. Los puntos de presencia fueron transformados en densidades de ocurrencia dentro del espacio ambiental definido por el PCA. A partir de estas densidades, se calculó el índice de ocupación, una medida que permite comparar de manera objetiva la distribución de la especie entre distintas regiones o condiciones ambientales, incluso cuando la disponibilidad de hábitats no es homogénea (Broennimann et al., 2014).
El índice de ocupación obtenido se representó en el espacio ambiental bidimensional para delimitar de manera precisa el nicho climático potencial de la especie. Posteriormente, se realizó un análisis de solapamiento y equivalencia de nicho, con el propósito de evaluar el grado de similitud y estabilidad. El procesamiento de las imágenes ráster se efectuó mediante el software ArcMap 10.2, con el fin de evaluar si las poblaciones de C. cateniformis ocupan condiciones ambientales similares a lo largo de su distribución geográfica o si existen diferencias ecológicas asociadas a gradientes climáticos o altitudinales. Este análisis permitió detectar zonas de convergencia y divergencia ambiental, proporcionando una comprensión más detallada de la estructura y extensión del nicho climático de la especie dentro del territorio amazónico ecuatoriano, contribuyendo así a identificar los factores ambientales que determinan su distribución potencial (Buendía-Espinoza et al., 2020).
Se realizó el posprocesamiento de los rásteres correspondientes al modelo de distribución potencial de C. cateniformis, los cuales fueron proyectados geográficamente en el software ArcMap 10.2 para su análisis espacial, visualización cartográfica e interpretación ecológica. El modelo generado produjo una superficie continua de probabilidad de presencia, cuyos valores se clasificaron en 5 categorías de idoneidad ambiental. Las áreas con valores inferiores al umbral se consideraron como ausencia potencial, mientras que los intervalos 0-0.25, 0.25-0.50, 0.50-0.75 y 0.75-1.00 fueron definidos como baja, media, alta y óptima probabilidad de presencia, respectivamente. Esta clasificación permitió una mejor discriminación de las zonas de mayor adecuación ambiental para la especie.
Los modelos finales se obtuvieron mediante una salida logística en el algoritmo MaxEnt, utilizando como criterios de desempeño la presencia mínima de entrenamiento y la maximización de la suma de sensibilidad y especificidad, lo que permitió establecer umbrales binarios de presencia-ausencia con alta confiabilidad (Scrivanti y Anton, 2020). Cada umbral fue estimado a partir del promedio de 100 iteraciones independientes a fin de reducir la variabilidad y garantizar la estabilidad del modelo. Finalmente, se identificaron y delimitaron áreas con alta idoneidad ambiental para el crecimiento y establecimiento de C. cateniformis en las 6 provincias de la región amazónica del Ecuador.
Resultados
Se obtuvo un valor de AUC = 0.992 para el modelado de C. cateniformis. Este valor corresponde a una alta capacidad de predicción y fiabilidad del modelo aplicado, lo que da confianza para reconocer la alta correspondencia entre las condiciones climáticas observadas y las áreas potenciales de distribución de la especie predicha por el modelo. La sumatoria de todas las aportaciones al modelo fue de 95.4%, donde la precipitación del mes más seco aporta 71.9%, seguido de la temperatura media del trimestre más seco (7.1%), la estacionalidad en la precipitación (3.4%), la precipitación del trimestre más frío (3%), la temperatura media anual (2.3%), la precipitación anual (1.8%), el rango de temperatura diurna (1.7%), la elevación (1.7%), la temperatura media del trimestre más lluvioso (1.5%) y, finalmente, la precipitación del trimestre más seco (1%).
Para la validación del modelo se utilizaron 136 registros independientes, conformados por 68 datos de presencia y 68 pseudoausencias generadas aleatoriamente dentro del área de estudio, evitando las zonas con registros confirmados de la especie. La matriz de confusión evidenció que el modelo clasificó correctamente 127 casos, de los cuales 64 correspondieron a presencias correctamente predichas (verdaderos positivos) y 63 a pseudoausencias correctamente identificadas (verdaderos negativos). Asimismo, se registraron 5 falsos positivos, donde el modelo predijo presencia en sitios clasificados como pseudoausencias y 4 falsos negativos, correspondientes a presencias observadas que no fueron detectadas por el modelo. Estos resultados evidenciaron una alta concordancia entre las predicciones y los datos de validación independientes, con una baja proporción de errores de comisión y omisión, lo que indicó un adecuado desempeño predictivo del modelo.
A partir de la matriz de confusión, se calculó el índice de Kappa y el estadístico de habilidad verdadera (TSS) para evaluar la precisión y la concordancia del modelo predictivo. El modelo alcanzó una sensibilidad de 0.94, evidenciando una alta capacidad para identificar correctamente las presencias reales de la especie y una especificidad de 0.93, lo que refleja una adecuada discriminación de las áreas no favorables. La combinación de estos parámetros derivó en un valor de TSS de 0.87, considerado indicativo de un desempeño óptimo y robusto, independiente de la prevalencia de la especie. De manera concordante, el índice de Kappa registró un valor de 0.87, lo que señala una alta coincidencia entre las observaciones y las predicciones del modelo, descontando el efecto del azar. La identificación de áreas con condiciones ambientales favorables para C. cateniformis se distribuyó principalmente en zonas con características ambientales específicas definidas por el modelo (fig. 2), lo que permitió delimitar espacialmente regiones con mayor idoneidad ecológica para la especie.
El PCA mostró que la precipitación del mes más seco (BIO14), explicó 71.90% de la contribución total al ajuste del modelo, constituyéndose como el predictor climático de mayor peso en la distribución potencial de C. cateniformis en la Amazonía ecuatoriana. Este resultado indica que la presencia de la especie se asocia principalmente a valores mínimos de precipitación mensual, mientras que áreas con déficits hídricos más marcados fueron clasificadas fuera del rango potencial de ocurrencia. La respuesta del modelo muestra una alta sensibilidad a las condiciones de precipitación mínima estacional y muestra que los valores bajos de BIO14 limitan la idoneidad ambiental para la especie. En consecuencia, el régimen hídrico mensual durante el periodo más seco se identificó como el factor climático dominante que condiciona la distribución espacial modelada de C. cateniformis.
La disponibilidad hídrica durante los periodos de déficit pluviométrico constituye el factor más restrictivo en la distribución potencial y sugiere una alta sensibilidad de la especie a las condiciones de estrés hídrico estacional. En menor proporción, la temperatura media del trimestre más seco (BIO 9) aportó 7.1%, mientras que la estacionalidad de la precipitación (BIO 15) y la precipitación del trimestre más frío (BIO 19) registraron contribuciones de 3.4 y 3%, respectivamente (tabla 2). Lo anterior sugiere que la variabilidad climática asociada a la disponibilidad de humedad y las condiciones térmicas durante los periodos críticos desempeñó un papel determinante en la configuración del nicho ecológico potencial.
| Código | Variables ambientales | Contribución (%) | Permutación (%) |
|---|---|---|---|
| BIO14 | Precipitación del mes más seco | 71.90 | 0.00 |
| BIO 9 | Temperatura media del trimestre más seco | 7.10 | 2.90 |
| BIO 15 | Estacionalidad en la precipitación | 3.40 | 34.90 |
| BIO 19 | Precipitación del trimestre más frío | 3.00 | 4.40 |
| BIO 1 | Temperatura media anual | 2.30 | 4.50 |
| BIO 12 | Precipitación anual | 1.80 | 0.10 |
| BIO 2 | Rango de temperaturas diurnas | 1.70 | 0.80 |
| Elevación | Elevación | 1.70 | 9.90 |
| BIO 8 | Temperatura media del trimestre más lluvioso | 1.50 | 8.00 |
| BIO 17 | Precipitación del trimestre más seco | 1.00 | 2.30 |
Las variables restantes presentaron contribuciones individuales menores a 2.5%, incluyendo la temperatura media anual (BIO 1) con 2.3%, la precipitación anual (BIO 12) con 1.8%, el rango de temperaturas diurnas (BIO 2) y la elevación con 1.7% cada una, así como la temperatura media del trimestre más lluvioso (BIO 8) con 1.5% y la precipitación del trimestre más seco (BIO 17) con 1%.
Aunque estas variables mostraron una influencia marginal, su incorporación permitió refinar la delimitación ambiental del modelo, aportando información complementaria sobre la variabilidad térmica y altitudinal del área de estudio. De manera agregada, las variables relacionadas con la precipitación acumularon más de 80% de la contribución total, lo que sugiere una fuerte dependencia de C. cateniformis de los gradientes de humedad característicos de la Amazonía ecuatoriana, destacando la importancia de los regímenes pluviométricos estacionales como principales moduladores de la distribución potencial de la especie.
Mediante el geoprocesamiento del SDM generado en el software MaxEnt, se determinó el potencial de distribución de C. cateniformis, que incluyó sitios definidos en las 6 provincias de la Amazonia ecuatoriana principalmente en la parte norte y sur (fig. 2). Específicamente las áreas óptimas para el desarrollo de esta especie se localizan en la provincia de Sucumbíos, al oeste del cantón Lago Agrio, norte y centro del cantón Cascales, al este del cantón Gonzalo Pizarro, al este del cantón Cuyabeno y finalmente en 70% del territorio perteneciente al cantón Shushufindi, mientras que en la provincia de Orellana la idoneidad óptima está al oeste del cantón Aguarico, este y oeste del cantón Puerto Francisco de Orellana, en 75% del territorio del cantón La Joya de los Sachas y al noreste del catón Loreto.
También hay condiciones óptimas en la provincia de Napo, al sureste del cantón Archidona, centro del cantón Tena y este del cantón Carlos Julio Arosemena Tola. De igual manera, en la provincia de Pastaza se registra una pequeña extensión al oeste del cantón Arajuno, mientras que en el cantón Santa Clara hay una pequeña extensión al oeste y en el cantón Pastaza. En la provincia de Morona Santiago las áreas con mayor idoneidad se encuentran al sur del cantón Taisha, gran parte del cantón Tiwintza y al sur del cantón Gualaquiza. Mientras que en la provincia de Zamora Chinchipe en gran parte del cantón Pangui, centro del cantón Yantzaza, oeste del cantón Paquisha, 80% del cantón Centinela del Condor, al este del cantón Zamora y, finalmente, al norte y este del cantón Nangaritza.
El área óptima del potencial de distribución de C. cateniformis fue de 13,288 km² (fig. 2). Esta superficie corresponde a la categoría de 0.75-1 para el desarrollo de la especie. En la Amazonía norte, conformada por las provincias de Sucumbíos y Orellana se concentra la mayor extensión de hábitat potencial, con 8,458 km²; en la Amazonía central, que comprende Napo y Pastaza, se identificaron 1,544 km²; y finalmente, en la Amazonía sur, correspondiente a Morona Santiago y Zamora Chinchipe, se estimaron 3,286 km². Considerando que la superficie total del ecosistema tropical en el ecuador es de aproximadamente 120,000 km², el área determinada por el modelo para C. cateniformis representa alrededor de 11.07% del territorio nacional con condiciones favorables para su supervivencia y producción (tabla 3), lo que evidencia un potencial significativo para su manejo sostenible en estas zonas.
| Ubicación | Provincia | Cantón | Km² | (%) |
|---|---|---|---|---|
| Amazonia Norte | Sucumbíos | Lago Agrio | 8,458 | 63.65 |
| Cascales | ||||
| Gonzalo Pizarro | ||||
| Cuyabeno | ||||
| Shushufindi | ||||
| Orellana | Aguarico | |||
| Puerto Francisco de Orellana | ||||
| La Joya de los Sachas | ||||
| Loreto | ||||
| Amazonia Centro | Napo | Archidona | 1,544 | 11.62 |
| Tena | ||||
| Carlos Julio Arosemena Tola | ||||
| Pastaza | Arajuno | |||
| Santa Clara | ||||
| Pastaza | ||||
| Amazonia Sur | Morona Santiago | Taisha | 3,286 | 24.73 |
| Tiwintza | ||||
| Gualaquiza | ||||
| Zamora Chinchipe | Pangui | |||
| Yantzaza | ||||
| Paquisha | ||||
| Centinela del Condor | ||||
| Zamora | ||||
| Nangaritza |
El modelo también identificó zonas con idoneidad moderada a alta en las provincias de Zamora Chinchipe y Morona Santiago, donde los patrones climáticos presentan una menor variabilidad anual y diaria. Estas áreas se caracterizan por una distribución más homogénea de la precipitación y una menor amplitud térmica diaria y estacional, lo que se refleja en la estabilidad espacial de la idoneidad predicha (fig. 3). Hay una diferenciación regional clara en las áreas con condiciones apropiadas para la especie dentro de la Amazonía ecuatoriana. Al superponer la distribución potencial obtenida en el modelo con las isotermas e isobaras presentadas en la figura 3, se observó que los sitios con mayor idoneidad se concentraron principalmente en la Amazonía norte y, de manera más marcada, en la Amazonía sur. Estos patrones espaciales coincidieron con las condiciones climáticas reportadas en los informes meteorológicos del Instituto Nacional de Meteorología e Hidrología (INAMHI, 2025), lo que permitió identificar regiones con mayor favorabilidad ambiental para C. cateniformis.
Los patrones climáticos en las provincias de Zamora Chinchipe y Morona Santiago muestran una mayor estabilidad tanto en el invierno como en el verano, así como en el día y la noche a lo largo del año; esta estabilidad se debe a la influencia de la cordillera Oriental que actúa como barrera orográfica que genera microclimas húmedos y regula la variabilidad térmica y pluviométrica. La topografía montañosa y la presencia de vastas áreas de bosques favorecen una distribución homogénea de la precipitación durante el año, con ciclos estacionales y diurnos menos marcados, lo que se traduce en una menor amplitud térmica diaria y estacional (fig. 3).
Las mayores probabilidades de idoneidad ambiental para C. cateniformis se asocian a rangos de precipitación entre 3,000 y 6,000 mm y a temperaturas comprendidas entre 22 y 26 °C. Estos rangos climáticos concentran los valores más altos de idoneidad predicha (fig. 3). En la Amazonía sur del Ecuador, particularmente en sectores influenciados por la cordillera Oriental, el modelo identificó áreas extensas y continuas con idoneidad alta, caracterizadas por condiciones de alta humedad y elevados niveles de precipitación. La distribución espacial de estas áreas muestra una relación consistente con gradientes altitudinales y climáticos propios de la región, evidenciando patrones regionales definidos de idoneidad ambiental para la especie.
Discusión
El modelado de nicho ecológico de C. cateniformis con 10 variables bioclimáticas de temperatura y precipitación ha permitido identificar áreas ambientalmente óptimas para el desarrollo y establecimiento de la especie dentro del bosque húmedo tropical, considerando sus requerimientos bioclimáticos. Los rangos de idoneidad ambiental obtenidos para C. cateniformis en la Amazonía ecuatoriana muestran coincidencias y diferencias con lo reportado por Rojas-Briceño et al. (2020) quienes registraron temperaturas entre 23 y 27 °C y precipitaciones de 1,800 a 2,440 mm. En este estudio, las mayores probabilidades de idoneidad se asociaron a temperaturas similares (22-26 °C), pero con precipitaciones más altas (3,000-6,000 mm). La similitud térmica sugiere que la temperatura representa un factor relativamente estable para la especie, mientras que la precipitación evidencia una mayor amplitud ecológica. Estas diferencias podrían estar relacionadas con las condiciones más húmedas características de la Amazonía ecuatoriana. En conjunto, los resultados sugieren que C. cateniformis presenta tolerancia a un amplio gradiente de precipitación, favoreciendo su distribución en ambientes altamente húmedos.
En la Amazonía alta del Ecuador, la topografía asociada a la cordillera oriental, frecuentemente referida como flanco oriental andino, modula de manera significativa la dinámica térmica y de humedad a escala local y regional (Arias-Gutiérrez et al., 2016). Las elevaciones montañosas actúan como barreras orográficas que interceptan los flujos de aire húmedo provenientes del Atlántico y favorecen procesos de ascenso forzado, condensación y precipitación recurrente (Vargas-Chaves, 2024). Este mecanismo incrementa la cobertura nubosa persistente y reduce la radiación solar directa durante el día, lo cual limita el calentamiento diurno, mientras que la alta humedad atmosférica y la nubosidad nocturna disminuyen la pérdida de calor por radiación, que atenúan el enfriamiento nocturno (Polasky et al., 2025). Como resultado, se generan microclimas con menor amplitud térmica diaria y estacional, así como una mayor estabilidad térmica a lo largo del año (Gomes-Oliveira et al., 2022). Estas condiciones, reforzadas por la compleja topografía y la continuidad de la cobertura boscosa, explican por qué las áreas identificadas por el modelo no experimentan cambios térmicos bruscos y mantienen niveles elevados de humedad, lo que contribuye a la persistencia de condiciones ambientales favorables para C. cateniformis.
En la Amazonía sur, Morona Santiago y Zamora Chinchipe se observó un foco de idoneidad notable que coincide con sectores donde la Cordillera Oriental genera una interacción orográfica que favorece microclimas locales con elevada humedad y ciclos convectivos frecuentes; esa influencia orográfica incrementa la retención de humedad y la frecuencia de eventos de precipitación intensa, favoreciendo sitios con condiciones climáticas óptimas para el establecimiento y crecimiento de C. cateniformis (INAMHI, 2025). Asimismo, los patrones estacionales ecuatorianos caracterizados por una estación lluviosa marcada por mayor transporte de humedad desde la cuenca amazónica y la circulación de vientos ondeantes, y por una estación relativamente más seca con mayor variabilidad interanual modulan la fenología y la disponibilidad hídrica en los microhábitats modelados. En este contexto, la interacción de masas de aire húmedo tropical transportada por la circulación atlántica y la convección local reforzada por sistemas convectivos y el reciclaje hídrico de la propia vegetación explican la persistencia de áreas con precipitación anual elevada 3,000-6,000 mm y temperaturas moderadas 22-26 °C que el modelo identifica como óptimos hallazgos concordantes con estudios recientes sobre climatología y dinámica convectiva en la Amazonía ecuatoriana (Calvas-Serrano et al., 2024).
Los patrones de distribución potencial obtenidos en este estudio muestran que las áreas con mayor idoneidad ambiental para C. cateniformis se concentran en zonas con alta precipitación y condiciones climáticas húmedas, lo que coincide con la dinámica climática de la Amazonia del Ecuador caracterizada por periodos lluviosos que favorecen el establecimiento de la especie (Vistin-Guamantaqui et al., 2025). Estas condiciones se relacionan con la influencia de masas de aire húmedo provenientes del Atlántico y la Amazonía, junto con el efecto orográfico de la cordillera, factores que contribuyen a generar los rangos de precipitación entre 3,000 y 6,000 mm identificados como óptimos en el modelo. Asimismo, las áreas con menor idoneidad coincidieron con zonas donde la estacionalidad seca es más marcada, lo que sugiere limitaciones en la expansión de la especie hacia ambientes marginales. Los resultados obtenidos refuerzan la importancia de las condiciones climáticas regionales y de la conservación de áreas con microclimas favorables para la persistencia y manejo sostenible de C. cateniformis en la Amazonía ecuatoriana.
De manera similar a lo reportado en la Amazonía norte del Perú, donde se proyecta un incremento de áreas con idoneidad media para C. cateniformis bajo escenarios futuros, los resultados obtenidos en la Amazonía ecuatoriana también evidencian la presencia de extensas zonas con idoneidad media distribuidas principalmente en las provincias de Pastaza, Morona Santiago, sectores de Orellana y áreas intermedias de Sucumbíos y Napo. Estas regiones, presentan condiciones subóptimas, con características ambientales favorables que podrían permitir la expansión potencial de la especie en el futuro (Polasky et al., 2025). Esta coincidencia entre ambos estudios sugiere que las áreas con idoneidad media representan zonas estratégicas para la planificación forestal y la implementación de sistemas agroforestales, especialmente en territorios que han sido alterados por actividades antrópicas. En este sentido, mientras el estudio realizado en Perú proyecta una expansión futura de estas áreas, los resultados obtenidos en Ecuador ya evidencian la existencia actual de estos espacios con potencial de establecimiento, lo que refuerza la importancia de considerarlos en estrategias de restauración, manejo sostenible y conservación de C. cateniformis a mediano y largo plazo (Gomes-Oliveira et al., 2022).
Por otro lado, es importante considerar lo señalado por Booth (2022), quien indica que la variabilidad climática interanual, especialmente aquella asociada a eventos como El Niño y La Niña, puede influir significativamente en estos ecosistemas. Esta situación resulta relevante, ya que C. cateniformis presenta una baja tolerancia a cambios bruscos en las condiciones ambientales. La alteración de los patrones de precipitación y temperatura durante estos eventos puede afectar su desarrollo en la cuenca amazónica y, en consecuencia, modificar o incluso reducir las áreas previamente identificadas como hábitats adecuados para la especie.
La implementación de sistemas agroforestales con C. cateniformis debe iniciarse mediante una planificación territorial basada en los resultados del modelo de distribución de especies (SDM), apoyada con herramientas SIG y visitas en campo para identificar áreas degradadas y suelos con potencial de recuperación ecológica dentro de los cantones antes determinados. Este proceso debe complementarse con la participación de las comunidades locales a través de talleres participativos, donde se integren conocimientos técnicos y saberes tradicionales para definir de manera conjunta los espacios más adecuados para el establecimiento de estos sistemas. Mediante representaciones cartográficas participativas, se pueden orientar las intervenciones hacia zonas estratégicas como áreas de recarga hídrica, corredores de conectividad ecológica y sitios con potencial de captura de carbono, favoreciendo la resiliencia de los paisajes agrícolas fragmentados frente al cambio climático. En este sentido, la articulación social y técnica, coinciden con lo señalado por Murniati et al. (2022), permite promover una gestión inclusiva, adaptativa y sostenible que fortalezca tanto la recuperación ecológica como los beneficios socioeconómicos para las comunidades locales.
En toda la Amazonía ecuatoriana se debe implementar un sistema de monitoreo ambiental participativo que trabaje conjuntamente con el Ministerio de Ambiente y Energía monitoreando parámetros como el almacenamiento de carbono, biodiversidad y calidad del suelo, las comunidades locales apoyadas por universidades pueden registrar el crecimiento de C. cateniformis la presencia de fauna asociada y la humedad del suelo mediante herramientas digitales, esto servirá para evaluar la efectividad de los sistemas agroforestales en la captura de carbono y la restauración ecológica. Se propone incluir C. cateniformis en programas nacionales de restauración productiva y mecanismos REDD+ (reducción de emisiones por deforestación y degradación forestal) orientados a la captura de carbono, donde a mediano y largo plazo el gobierno podría ofrecer incentivos tributarios o económicos a productores que adopten modelos agroforestales sustentables. Además, se recomienda integrar los resultados del SDM en los planes de ordenamiento territorial provincial y garantizar así la compatibilidad entre desarrollo rural y conservación. Esta articulación institucional fomentará la resiliencia ecológica, la recuperación de la biodiversidad y el cumplimiento de compromisos climáticos internacionales del Ecuador.
Agradecimientos
Los autores expresan su agradecimiento a la Escuela Superior Politécnica de Chimborazo (ESPOCH) por el apoyo institucional brindado para el desarrollo de la presente investigación.
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The reef-associated fishes of the Revillagigedo National Park (Mexican Pacific): a curated, annotated and illustrated inventory
Los peces asociados a arrecifes del Parque Nacional Revillagigedo (Pacífico mexicano): un inventario curado, comentado e ilustrado
D. Ross Robertson a, *, Benjamin Frable b, Benjamin C. Victor c, William Ludt d, Michelle R. Gaither e, Carlos J. Estape f, Allison Morgan-Estape f, Diane Pitassy g, Katriina L. Ilves h, Omar Domínguez-Domínguez i, Omar Valencia-Méndez j, Arturo Ayala-Bocos k, Edgar Adrián Acevedo-Álvarez i, María Gloria Solís-Guzmán i, A. Karim Awhida-Robinson i, Luis Fernando Martínez-García i, Ayla Gabriela Ramos-Mayoral i, María Elena Castillo-Victor i, Xavier Madrigal-Guridi i, Juan Esteban Martínez-Gómez l, Eloísa Torres-Hernández m, Lee Richter n, Sara Richter n, Andre Michael Goolishian-Hernández o, Serenity Mitchell o, Jeffrey D. Haines p, Alasdair G. Dunlap-Smith q, Ellen Place r, Carol Cox s, Robert Cox s, Fernando Duarte-Ramos t, Carlos Armando Sánchez-Ortiz t
a Smithsonian Tropical Research Institute, Naos Marine Laboratory, Ft Amador, Balboa, Republic of Panamá
b University of California San Diego, Scripps Institution of Oceanography, Marine Vertebrate Collection, 9500 Gilman Drive, La Jolla, CA 92093-0244, USA
c Ocean Science Foundation, 4051, Glenwood, CA 92604, USA
d Natural History Museum of Los Angeles County, Department of Ichthyology, 900 Exposition Blvd., Los Angeles, CA 90007, USA
e University of Central Florida, Genomics and Bioinformatics Cluster, Department of Biology, 4110 Libra Drive, Orlando, FL 32816, USA
f 197 Gulfview Drive, 33036 Islamorada, FL, USA
g Smithsonian Institution, National Museum of Natural History, Museum Support Center, 4210 Silver Hill Rd., Suitland, MD 20746, USA
h Canadian Museum of Nature, Research & Collections, P.O. Box 3443, Sta. D, Ottawa, Ontario, K1P 6P4 Canada
i Universidad Michoacana de San Nicolás de Hidalgo, Facultad de Biología, Laboratorio de Biología Acuática, Edificio R, Planta Baja, Ciudad Universitaria, 58030 Morelia, Michoacán, Mexico
j Centro de Investigación Científica y de Educación Superior de Ensenada, Departamento de Ecología Marina, Carretera Ensenada – Tijuana No. 3918, Zona Playitas, 22860 Ensenada, Baja California, Mexico
k Ecosistemas y Conservación, Proazul Terrestre A.C., Benito Juárez 665, Centro, 2300 La Paz, Baja California Sur, Mexico
l Instituto de Ecología A.C., Red de Interacciones Multitróficas, Carretera antigua a Coatepec 351, Col. El Haya, 91073 Xalapa, Veracruz, Mexico
m Universidad Nacional Autónoma de México, Instituto de Biología, Departamento de Zoología, Pabellón Nacional de la Biodiversidad, Colección Nacional de Peces, Tercer Circuito, Ciudad Universitaria, Coyoacán, 04510 Ciudad de México, Mexico
n Fish Bay #1, 00830 St. John VI, USA
o 5000 estate enighed, PMB 337, 00830 St. John USVI, USA
p 2310 SE Maniton Terrace, 34952 Port St. Lucie, FL, USA
q 50 Tug Mountain Road, Orange, 03741 New Hampshire, USA
r 17117 NE 5th St. 98008 Bellevue, WA, USA
s 202 Coral Dr, 32456 Port St. Joe, FL, 32456 USA
t Universidad Autónoma de Baja California Sur, Departamento de Ciencias Marinas y Costeras, Carretera al Sur Km. 5.5, 23080 La Paz, Baja California Sur, Mexico
*Corresponding author: robertsondr@si.edu (D.R. Robertson)
Received: 27 August 2025; accepted: 13 April 2026
Abstract
Accurate inventories of species living at isolated islands are essential for understanding the biogeographic and functional composition of insular faunas, habitat area vs. faunal-size relations and levels of endemism, as well as for purposes of conservation and management. Constructing such inventories involves new field research and rigorous winnowing of existing records that have errors in specimen identification and metadata or are unsubstantiable due to a lack of reliable source information. This paper provides a curated, photographically illustrated inventory of reef-associated fishes found at the 4 islands in the Revillagigedo National Park (RNP), 400 km southwest of Baja California. We assessed the quality of data about records relating to 364 species and added new data (specimens of 132 species and images of 182 species) that confirm species occurrences at each island in the park. The resultant inventory includes 234 accepted species occurrences, with 121 unaccepted occurrences (33.2% of the assessed species). Among the confirmed species, 166-189 (70.9-80.8%) evidently are park residents. New information indicates that the RNP fauna includes 23 named and probable species of endemics, representing 9.8% of the accepted species and 12.2-13.9% of the residents, highlighting the uniqueness of the RNP and its importance for conservation.
Keywords: Tropical Eastern Pacific; Isolated oceanic island; Rocky reefs; Elasmobranchs; Bony fishes
Resumen
Los inventarios precisos de especies en islas aisladas son esenciales para comprender la composición biogeográfica y funcional de las faunas insulares, las relaciones área-fauna y los niveles de endemismo, además de ser fundamentales para la conservación y manejo. Su elaboración requiere nuevas investigaciones de campo, además de un riguroso cribado de registros existentes con errores de identificación o metadatos no verificables. Este artículo presenta un inventario ilustrado de los peces asociados con arrecifes del Parque Nacional Revillagigedo (PNR), ubicado a 400 km al suroeste de Baja California. Se evaluó la calidad de registros de 364 especies, incorporando datos nuevos (especímenes de 132 especies e imágenes de 182) que confirman presencias en cada una de las islas. El inventario resultante incluye 234 presencias aceptadas y 121 no aceptadas (33.2 % de las especies evaluadas). Dentro las confirmadas, entre 166 y 189 especies (70.9%-80.8%) son residentes del parque. La nueva información indica que la fauna del PNR incluye 23 especies endémicas nombradas y probables, representando 9.8% de las especies aceptadas y entre 12.2% y 13.9% de los residentes, lo que ilustra la singularidad ecológica del PNR y su importancia para la conservación marina.
Palabras clave: Pacífico oriental tropical; Isla oceánica aislada; Arrecifes rocosos; Elasmobranquios; Peces óseos
Introduction
Accuracy in faunal inventories for reef fishes and other reef organisms at isolated tropical islands is important for a variety of reasons. Species inventories often provide the only data used to assess the biogeographic relationships Dubic et al., 2023; Hobbs et al., 2012, 2014; Robertson & Allen, 1996; Robertson & Cramer, 2009), as well as many studies of their composition in terms of functional ecological characteristics (e.g., Bender et al., 2013; Dubuc et al., 2023; Ferrari et al., 2023; Palacios-Salgado et al., 2019; Torres-García et al., 2025), relationships between faunal size, habitat area and island isolation (Dubuc et al., 2024; Hobbs, 2012, 2014; Olivier et al., 2018; Sandin et al., 2008) and levels of endemism (Dubuc et al., 2023; Hobbs et al., 2011, 2014; Robertson & Cramer, 2009; Victor, Grove et al., 2024). Accurate faunal inventories are also essential for management and conservation purposes (IOC-UNESCO, 2024).
Faunal inventories typically are assembled based on various combinations of information in existing scientific and grey-literature publications and in museum databases, some hosted by online aggregators, that refer to specimens collected from the 18th century to the present. They also often include personal observations by authors and the results of new research. However, as time goes by, inventories tend to accumulate species names through the uncritical acceptance of records. Constructors of inventories often fail to question and delete names except those representing obvious errors, such as the inclusion of species that do not occur in the biogeographic region in which the study area is located. However, data from primary sources of information, such as museum collections, particularly older collections, also need to be rigorously evaluated to ensure accuracy in species occurrences on inventories. For example, Mundy (2005) compiled a curated inventory of Hawaiian Islands fishes, which involved a detailed annotated evaluation of supporting information for each species. That inventory included 1,177 verified marine species plus 9.7% of the total evaluated that he regarded as invalid. Another inshore-fishes inventory by Mundy et al. (2010) for parts of the Phoenix and Line Islands included 5.6% of the species as unaccepted or requiring verification. Rates of rejection in other curated local marine fish inventories in some cases are even higher: 14.4% of marine fishes recorded at Malta (Borg et al., 2023) and 18.1% of alien marine fishes recorded at Greece (Zenetos, 2018).
The Revillagigedo Archipelago is an isolated cluster of 4 small, volcanic islands that lie 400 km southwest of the tip of the Baja California Peninsula, Mexico, in what is now the Revillagigedo National Park (RNP), which receives the highest level of protection available to Mexican parks, with no industrial fishing allowed. Research on the archipelago’s fish fauna began with collections between the 1880s and 1890s studied by Gilbert (1890), Heller and Snodgrass (1903), Jordan and Gilbert (1882), Jordan and McGregor (1899), and Snodgrass and Heller (1905). This was followed by another burst of collecting activity in the mid-20th century (Beebe, 1937; Ricker, 1959; Schmitt & Schultz, 1940) and subsequent deposition of substantial collections of specimens made between the 1930s and 1970s at the California Academy of Sciences, the Natural History Museum of Los Angeles County, the University of British Columbia Fish Collection (now part of the Beaty Biodiversity Museum) and the Canadian Museum of Nature, both in Canada, Scripps Institution of Oceanography and the U. S. National Museum of Natural History. In Mexico there are collections of RNP fishes at the Colección Nacional de Peces (CNPE-IBUNAM), the Colección del Laboratorio de Ecología, Escuela Nacional de Ciencias Biológicas, Instituto Politécnico Nacional (ENCB-IPN-LEM), the fish collection at Universidad Michoacana de San Nicolás de Hidalgo (CPUM) and the Colección Ictiológica, Centro Interdisciplinario de Ciencias Marinas, Instituto Politécnico Nacional, Baja California Sur (CICIMAR).
The first major ichthyological inventory of RNP fishes that incorporated information from a variety of sources was that of Castro-Aguirre and Balart (2002). That inventory, which focused on the biogeographic relationships of members of that fauna, included 228 shore —and pelagic fishes— 178 of them reef-associated species. However, it lacked original (primary) source records linked to each of the species included and did not indicate occurrence of each species at the different RNP islands. Subsequently, 2 more comprehensive inventories of RNP fishes were published in 2016. Both included reef-associated and other types of fishes, provided extensive primary-source records and employed some degree of winnowing of doubtful and erroneous records. One inventory, by Del Moral-Flores et al. (2016), which provided information on species occurrences at each of the 4 RNP islands, excluded 36 (8.9%) of the total of 402 species it discussed, most due to changes to nomenclature in the taxonomic literature, but some arising from issues with the geographic ranges of certain species and possible inappropriateness of archipelago habitats for others. The other inventory, by Fourriére et al. (2016), which did not include information on occurrences at each of the RNP islands, classed 10 (2.5%) of 399 species as doubtful occurrences.
Some of the RNP specimens from 19th-mid 20th century museum collections have been reassessed during taxonomic studies that targeted specific groups of archipelago fishes. These include species in various families of blennies (Dawson, 1975; Hastings & Robertson, 1999; Hubbs, 1953; Kresja, 1960; Rosenblatt et al., 2013; Springer, 1962); wrasses (Allen & Robertson, 1992a; Bussing, 1987); pipefishes (Dawson, 1981, 1985; Fritsche, 1980); gobies (Bussing, 1990; Ginsburg, 1947; Miller & Stefanni, 2001); soapfishes and groupers (Allen & Robertson, 1999; Guimares, 1999; McCarthy, 1979); damselfishes (Allen & Woods, 1980); clingfishes (Briggs, 1951, 1955); needlefishes (Collette & Banford, 2001); and frogfishes (Pietsch & Arnold, 2020). However, a substantial fraction of the museum material and its provenance have yet to be reexamined in the light of current taxonomic knowledge. That knowledge, now readily accessible in vast online resources, was not available to people responsible for making and annotating 19th to mid-20th century faunal collections and documenting that information in publications.
Here we expand on the “two 2016 efforts” to produce an improved inventory based on a more comprehensive verification of species records. In addition, in November 2022 a group of 17 scuba divers (7 diving ichthyologists plus 10 citizen-scientist underwater photographers) from Mexico, the USA and Panama spent 9 days diving at the 4 islands of the RNP. The objective of that expedition was to document occurrences of reef fishes at each island by photographing as many species as possible in their natural habitats and collecting specimens of some for further study (see Estape [2023] for a video presentation of that expedition). The present paper developed from that effort by including detailed assessments of species in the RNP based on various types of pre-existing information to produce a comprehensively curated inventory of the archipelagos reef-associated fishes that also documents their distribution among the different islands. That pre- existing information includes records in professional ichthyological publications that provide useable primary- source information: accessible and assessable species-level location data and selected museum specimens that date from collections in the later 19th to mid-20th century that have never been reassessed since those specimens were accessioned by the museums. This paper also incorporates a substantial amount of additional new primary-source data obtained in the RNP at different times during the past decade by different coauthors: specimen collections, DNA analyses, images of live and freshly collected specimens, and reliable field observations.
Materials and methods
The Revillagigedo Archipelago comprises 4 small volcanic islands situated southwest of the tip of the Baja California peninsula, with the closest island, San Benedicto, 406 km from the tip of Baja and the farthest, Clarion, 700 km from Baja. Three of the islands (Socorro, San Benedicto, and Roca Partida) are relatively close together (50-140 km apart) on the eastern side of the archipelago and are well separated from Clarion Island, the westernmost island, by distances ranging from 290 to 430 km. These islands are the base of the Revillagigedo National Park (the RNP), a rectangular 148,000 km2 marine park bounded by latitudes 17.655° to 20.009° N and longitudes -110.078° to -115.471° W. That reserve encompasses the Revillagigedo archipelago and a surrounding area of 6,400 km2 of deep sea, which is imbedded in a 142,000 km2 buffer zone (Fig. 1). The 4 islands are relatively small, ranging from < 1 ha. for Roca Partida to ~ 130 km2 for Socorro. Marine habitats consist mainly of rocky shorelines plus some soft bottoms, which drop rapidly off into deep water, intertidal rock pools and a few sheltered bays on Socorro and Clarion that have sandy beaches, with 1 on Socorro having a marine lagoon behind it. Reefs of the RNP islands are predominately rocky (Ramírez-Zúñiga et al., 2025), although scattered coral colonies and communities are common.
Here we follow the family and genus level classification of fishes as indicated in Eschmeyer’s Catalog of Fishes (Fricke et al., 2026).
Reef-associated fishes
Coral and rocky reefs down to depths of ~ 250 m have reef-fish faunas dominated by members of families of bony fishes found associated with shallow reefs (Baldwin et al., 2018). The inventory presented here focuses on species belonging to such families because those are what are traditionally considered to be reef fishes. Reef-associated fishes include not only demersal and benthic species that live in or on consolidated hard substrata (coral- and rocky reefs) but also demersal and benthic species that are restricted to unconsolidated or soft bottoms (sand, mud, and gravel) immediately adjacent to or within the matrices of reefs. Reef-associated fishes also include pelagic species that live in the water column, facultatively associate with reefs, are regularly seen over and immediately adjacent to them, and have trophic interactions with organisms on reefs. Those trophic interactions can include contributing food to reefs (e.g., material removed from the skin of large pelagics by reef fishes, e.g., Mobula birostris in the RNP (Clarión Angelfish & Manta) or Mola alexandrini at the western Galápagos (Bodianus & Mola) or extracting food from them (e.g., Carcharhinus falciformis and Thunnus albacares preying on juvenile Elagatis bipinnulata on reef at Cocos Is.; Auster et al., 2016). This definition follows Cord et al. (2024) and Robertson (2024).
The inventory presented here, which includes both bony fishes and elasmobranchs, is based on a wider range of species than those classed as reef fishes by Fourriére et al. (2016), who split the fauna into reef fishes and non- reef fishes, with the latter group including many pelagic and soft-bottom fishes that we class as reef-associated (see Robertson [2024] for a list of Tropical Eastern Pacific (TEP) reef-associated bony-fish species), as well as deep-water species belonging to non-reef families. The inventory of Del Moral-Flores et al. (2016) did not consider reef-associated fishes as a separate group.
Categories of species evaluated
The categories are: i) Confirmed/accepted species: species confirmed as part of the RNP fauna comprise those with positive identifications from 1 or more primary sources (see below). Accepted species also include a few common, widely distributed TEP species with distinctive morphological features that make them readily recognizable to divers and that were observed at 1 or more islands by a coauthor of this paper or in existing published professional ichthyological studies. They also include settlement-stage juveniles of benthic and demersal reef-associated species that were collected at a night-light while the ship was anchored immediately adjacent to the 3 largest islands during the 2022 expedition. ii) Unaccepted species: such species lack any identifiable support from primary sources that would confirm their presence in the RNP. They include misidentified museum specimens reassessed as such by museum curators, museum specimens that are too small or in such poor condition as to be unidentifiable to species, those with inadequate or erroneous location data, those whose occurrence is based on museum specimens that have been all lost and demersal/benthic species recorded as present based on specimens of pelagic larvae collected well offshore rather than benthic or settlement stage individuals. iii) Unresolved species: these include species that have been recorded at and plausibly could occur in the RNP, based on their geographic ranges, and occurrence at other TEP offshore islands, but currently lack adequate primary source data that verify their records. They also include species recorded as observed in previous studies that might have been confused with other visually similar species confirmed as present in the RNP.
Residency status
We regard a reef-associated species as being a resident in the RNP if there is evidence of long-term, persistent occurrence or an extant breeding population at 1 or more of the 4 islands. That evidence includes multiple observations of adults and, especially, juveniles at 1 or more of those islands during the 2022 expedition, repeated collections of multiple individuals over many years, multiple vouchered records at 1 or more islands over multiple recent years of ecologically non-cryptic species and recent images from iNaturalist and other sources at those islands. We used the yardstick of “multiple occurrences at multiple islands and/or multiple, well separated years, particularly if there are records since 2000”. This leaves a group we class as uncertain residents. These mainly include a few species that have a single, usually old record in the RNP and that lack recent records due to sampling difficulties may have led to their abundance being underestimated. Among them are some cryptic species, such as ophichthid eels and flatfish that bury in soft bottoms or live deep inside crevices (bythitids) that are normally effectively sampled only with ichthyocides. They also include species typically found below normal scuba depths (e.g., some congrids, tilefishes, serranids, scorpionfishes, and bythitids), that are less likely to have been sampled. Some may eventually be shown to be residents, while others may be vagrants. Non-residents or vagrants: species that do not fit either of the above categories and that have sufficiently few records that they likely represent species that occasionally recruit to the RNP but do not establish a breeding population there. They mainly include demersal, readily visible species living in shallow water, that, if they were residents, would be expected to be recorded repeatedly, in substantial numbers, over multiple years.
Endemics and the rate of endemism
Endemics are species that have a resident population only in the RNP. They include some species known from the mainland or other islands in small enough numbers to lack self-sustaining populations outside the RNP. Genetic evidence of endemism includes RNP populations composed entirely of local haplogroups or almost entirely when there are vagrant individuals at other sites, while non-endemics have substantial amounts of shared haplogroups with other sites, on the mainland and other TEP oceanic islands.
Levels of endemism frequently are measured as the percentage of the total number of species known from a location that are endemics, because those are the only data available. Examples for TEP fishes include Briggs (1974), Briggs and Bowen (2013), Del Moral-Flores et al. (2016), and Fourriére et al. (2014, 2016, 2017). In their discussion of geographic patterns of endemism among shore-fishes of the TEP, Robertson and Cramer (2009) considered endemism in relation to numbers of regional residents and excluded regional vagrants (c.f., Pinheiro et al., 2018). However, they did not consider more localized (provincial or island) endemism in relation to site residency and used the percentage of all species present at each site because residency information was not available for most oceanic islands. McCosker and Rosenblatt (2010) assessed the rate of endemism among Galápagos fishes in relation to local (Galápagos) residency, excluding local vagrants, while Victor, Grove et al. (2024) assessed it both in relation to the total fauna and to residents. If endemism is related to the total local fauna its rate will gradually decline through time as new vagrants appear at a site, even though the number of resident species may remain constant. Vagrants also have quantitatively negligible and temporary influence within local ecosystems, on resident populations or on the evolution of endemic populations. Hence, here we also focus on residents in calculating the rate of endemism for RNP reef-associated fishes.
Data sources
Primary sources used here comprise various types of sources that provide species-level data relating to occurrence of fishes within the RNP that we were able to assess for validity. They include information on specimen records in museum databases and online aggregators of such data, scientific publications by professional ichthyologists that recorded species during fieldwork within the RNP and taxonomic publications that involved examination of museum specimens from the RNP. They also include taxonomically diagnostic images of species taken in the RNP, including images taken by the authors and by visitors to the archipelago that are hosted by iNaturalist. Those sources also include DNA sequence data obtained from collections in the archipelago by some of this paper’s coauthors. The present paper is based on information from primary sources, which include 2 previous comprehensive inventories, by Del Moral-Flores et al. (2016) and Fourriére et al. (2016) — collectively referred to as the “two 2016 inventories”— both of which listed museum data for individual species. Del Moral-Flores et al. (2016) also listed which islands each species had records from, while Fourriére et al. (2016) recorded observations of species in different years, without indicating species occurrences at individual islands in the archipelago. Del Moral-Flores et al. (2016) referred to collections from 4 Mexican museums and 11 museums in other countries and mentioned the physical review of museum specimens. However, besides including a figure of images of 10 common species from 2 museum collections, the Colección Nacional de Peces, IBUNAM (CNPE-IBUNAM) in Mexico and the Natural History Museum of Los Angeles County (LACM) in the USA, that paper does not indicate which other specimens from which museums were examined for verification. Fourriére et al. (2016) reported no examination of any museum specimens.
We used several types of data to vouch for (i.e., validate) the occurrence of different species at different islands in the RNP. i) Taxonomically diagnostic images and specimens obtained during the 2022 expedition. The November 2022 expedition visited all 4 islands on a liveaboard diving support ship, the Quino El Guardián (see Estape, 2023). ii) Taxonomically diagnostic iNaturalist images. Images of fishes taken by citizen scientists available on the iNaturalist website (iNaturalist) that we reviewed and confirmed the identification of provided voucher information about occurrences of species at different islands in different years. iii) Universidad Michoacana Team collecting efforts between 2015-2023 and genetic analyses of RNP fish populations. Groups of ichthyological students from the Universidad Michoacana (UMSNH) laboratory of coauthor ODD made 4 collecting trips (2015, 2016, 2019 and 2023) to Clarion, San Benedicto and Socorro, collecting as many species as possible, particularly those found in tide pools. That collecting by ODD’s laboratory has led to genetic evaluations of the relationships of RNP populations to those on the mainland and other oceanic islands in the TEP, information that we draw on here. In addition, coauthor AAB made a series of visits to various parts of the archipelago in between 2012, 2013, 2014, 2017, 2019, 2022, 2023, 2024 and 2025, where he recorded species he observed while diving at different islands and, in some cases, took underwater videos of unusual species, from which screen captures provide island-specific voucher images used here. iv) UABCS, La Paz efforts between 2016 and 2025. Coauthor CS-O participated in an expedition to the archipelago during 2016 that resulted in a technical report (Aburto-Oropeza et al., 2016) based largely on several thousand hours of videos recorded by BRUVS (Baited Remote Underwater Video Systems). That report provided voucher data for a review of the occurrences of elasmobranchs and chimaeras in the RNP by Becerril-Garcia et al. (2020), which we regard as a primary-source curated inventory. Those videos also produced voucher images of deep- living bony reef fishes species included here. Monitoring reef-fish populations in the RNP by CS-O’s UABCS-La Paz team also produced the most recent record of a new member of the RNP reef-associated fish fauna, in June 2025. v) Reassessment of museum specimens. These were made in cases where no other primary sources of data confirmed the presence of a species in the RNP. In such cases fish curators at various museums examined specimens of particular species to verify their identities. Associated collection-site metadata was also reviewed to verify whether they were collected in the RNP. That information came from a combination of aggregators, including GBIF, OBIS, Conabio (Comisión Nacional para el Conocimiento y Uso de la Biodiversidad), Fishnet2, Vertnet, and iDigBio. In addition, the online databases of museums that supplied data to those aggregators were reviewed because no aggregator provides all the data currently available in individual museum databases that feed data to that aggregator. Museum name codes follow Sabaj (2023). Those museums are: ANSP (Academy of Natural Sciences, Philadelphia), AMNH (American Museum of Natural History), CAS (California Academy of Sciences, San Francisco), CMNFI (the Canadian Museum of Nature), CNPE (Colección Nacional de Peces, IBUNAM, Mexico), CPUM (Colección de Peces de la Universidad Michoacana de San Nicolás de Hidalgo, Morelia), CSLB (California State University, Long Beach, Fish specimens), CUMV (Cornell University Museum of Vertebrates, Ithaca), FMNH (Field Museum of Natural History, Chicago, Zoology Fish Collection), LACM (Natural History Museum of Los Angeles County), MCZ (Museum of Comparative Zoology, Cambridge), NHMUK (Natural History Museum, London), ROM (Royal Ontario Museum, Toronto), SIO (Scripps Institution of Oceanography, San Diego), UBC (University of British Columbia, Beaty Biodiversity Museum; which has too many errors to represent a very reliable source), UF (University of Florida, Florida Museum of Natural History, Ichthyology), UMMZ (University of Michigan Museum of Zoology), USNM (National Museum of Natural History, Smithsonian Institution, Washington DC). In cases where questionable specimens of a particular species were lodged at multiple museums, specimens from only 1 museum typically were reviewed. If the identity in that database was confirmed, we assumed correct identifications of specimens at any other museums, since the object of these reviews was to verify occurrence in the RNP. In situations in which misidentifications or erroneous location data were detected in the first case reviewed, specimens and collection data from another museum were examined, if available. vi) IATTC geographic distribution data of species caught in the international tuna fishery. We also used public domain location records of pelagic species caught in the TEP by the international tuna fishery and recorded by observers of the Inter-American Tropical Tuna Commission (IATTC) to help determine which of those species that are reef-associated are found in the RNP. vii) Genetic analyses of 2022 expedition specimens. We used these to confirm identifications of certain species and to provide data on evolutionary relationships of RNP populations with those on other TEP oceanic islands and the mainland. viii) UNESCO eDNA data. In 2023 a UNESCO sponsored eDNA sampling of water at a total of 21 stations adjacent to all 4 islands. The data on species-level assignments and location of ASV reads of 3 mitochondrial genes, cytochrome oxidase-1 (CO1), 12S and 16S for each species are located at a UNESCO site (UNESCO eDNA) and at OBIS (OBIS-UNESCO eDNA). We reviewed those data and used them to add a few species to the RNP confirmation-required group of-species that had not previously been recorded there and expanded the base of voucher data on some other species that had other existing forms of voucher data. Three species that lack other forms of voucher data have CO1 species-name assignments at the 100% confidence level, as well, in some cases, 16S reads with name assignments with 100% confidence. Given problems in name assignment due to misidentification of reference sequences in GenBank and limitations on the ability of 12S and 16S to reliably distinguish between congeners (Fontes et al., 2024), we added only species in cases in which misidentifications or confusion with other members of the same genus are unlikely, e.g., monospecific families or genera, when a species was the sole member of its genus in the TEP or when all other congeners present in the TEP were also included in the eDNA database. Other eDNA data, again mostly CO1, with 100% confidence species-name assignments, provided additional information on the occurrence and distribution in the RNP of 14 other species already confirmed by other voucher data. ix) Video surveys of deep-living fishes. There have been a few studies of deep-reef fishes in the RNP fauna based on videos collected by dropcams (Giddens et al., 2019), a small ROV (Remotely Operated Vehicle) operated at 30-90 m depth of both reef and adjacent sand habitat (Hollarsmith et al., 2020; Velasco-Lozano et al., 2020) and a submersible (Ayala-Bocos et al., 2015). These have provided useful information through voucher images of identifiable species at different RNP islands and information on mesophotic reef-fish abundances. However, much of the video databases produced by Aburto-Oropeza et al. (2016) and Ayala-Bocos et al. (2015) have yet to be reviewed and may in the future produce confirmation of additional species not mentioned here, as well as information that could help clarify the population status of some species already confirmed from the RNP.
Secondary sources are those that do not provide species-level data linked directly to a primary source, e.g., a georeferenced location from a museum record, or an image or indication from which island(s) in the RNP a species was recorded. Secondary sources include grey literature and government reports about the RNP that include lists of fish species. They also include some ichthyological publications that produce species lists based on data recycled from other published sources and do not individually link listed species to primary sources or to particular islands, e.g., Castro-Aguirre and Balart (2002) and Conanp (2019) and the governmental reports derived from those. The RNP species list used by Torres- García et al. (2025) to examine variation in taxonomic and functional diversity of fishes among various Mexican Pacific reefs also does not include primary-source data, does not refer to either the 2002 or 2016 checklists and appears to have used lists from government reports that recycle secondary data. Hence this paper is treated as an unusable secondary source. Other secondary sources also include digital resources by Robertson and Allen (2006, 2015) for which no primary-source information could be found in 2025 on some RNP location records. Hence, such secondary data cannot be reassessed if the basis for inclusion of each species on an inventory is unclear and such data are not relied upon here.
Results
Accepted species
We assessed the quality of existing primary-source records relating to 364 species included in various RNP databases and added new data (specimens of 132 species and images of 182 species) that confirm species occurrences at the different islands and assist estimation of their population status. Acceptance was based on a variety of factors. While reliable images demonstrated occurrences of many species, most also had other forms of supporting data. Re-examination of specimens of 46 species from 8 museums confirmed their identifications and location data, including 30 species lacking other forms of voucher data. Re-examination of museum specimens also corrected misidentifications of specimens of 12 other species, providing additional data relating to their status as RNP residents (Table 2). The resultant inventory includes 234 accepted species occurrences (including 19 not in both 2016 inventories and another 5 not in one of them), from 165 (80.5%) of the 205 assessed genera and 66 (81.5%) of the 81 assessed families (Tables 1, Supplementary material: Table S1). Among the confirmed species, 166-189 (70.9-80.8%) have sufficient records to be regarded as archipelago residents. That proportion may well increase when more information becomes available on non-resident and “residency uncertain” species.
| Family | Resident | Species | Clarion | Roca Partida | San Benedicto | Socorro | Fourr. accepted | DMF accepted | Supporting data |
|---|---|---|---|---|---|---|---|---|---|
| Acanthuridae | yes | Acanthurus nigricans | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); inat; observe; 2022- images, specimens, common |
| Acanthuridae | yes | Acanthurus triostegus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (9); inat; observe; 2022- images, specimens, schools |
| Acanthuridae | yes | Acanthurus xanthopterus | Yes | Yes | Yes | Yes | Yes | Yes | Museum (1); inat; observe; 2022- images, singletons |
| Acanthuridae | uncertain | Ctenochaetus marginatus | Yes | Yes | Yes | Yes | Yes | Museum (1); inat; observe; 2022- images, singletons | |
| Acanthuridae | yes | Prionurus laticlavius | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); inat; pubs; 2022- specimens, images, common |
| Alosidae | no | Sardinops sagax | Yes | (Yes) | Yes | Museum* (1) | |||
| Antennariidae | yes | Abantennarius sanguineus | Yes | Yes | Yes | Yes | Yes | Museums (4); eDNA; 2022- images, specimens, DNA | |
| Antennariidae | no | Antennarius commerson | Yes | Yes | Yes | 2022- specimen, image | |||
| Antennariidae | no | Antennatus strigatus | Yes | Yes | Yes | Museum* (1), image | |||
| Antennariidae | yes | Fowlerichthys avalonis | Yes | Yes | Museum (1); pub; images | ||||
| Anthiadidae | yes | Pronotogrammus multifasciatus | Yes | Yes | Yes | Yes | Museum (1); eDNA | ||
| Apogonidae | yes | Apogon atricaudus | Yes | Yes | Yes | Yes | Yes | Museums (6); inat; pubs; 2022- specimens, images, common | |
| Apogonidae | no | Apogon retrosella | Yes | Yes | Yes | 2022- specimen, image, DNA | |||
| Atherinopsidae | yes | Atherinella eriarcha | Yes | Yes | Yes | Yes | Yes | Museums (4); inat; pubs; 2022- specimens, DNA, common | |
| Aulostomidae | yes | Aulostomus chinensis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); inat; observe; 2022- images |
| Balistidae | yes | Balistes polylepis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); inat; pubs; observe; 2022- images |
| Balistidae | uncertain | Canthidermis maculata# | Yes | Yes | Yes | Yes | Museum* (2); pub | ||
| Balistidae | yes | Melichthys niger | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); pubs; inat; observe; 2022- images, specimens, common |
| Balistidae | yes | Sufflamen verres# | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (7); pubs; inat; observe; 2022- images, specimens, common |
| Balistidae | yes | Xanthichthys mento | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (8); pubs; inat; observe; 2022- images, specimens, common |
| Belonidae | yes | Ablennes hians# | Yes | Yes | Yes | (Yes) | Yes | Observe; eDNA | |
| Belonidae | yes | Platybelone pterura | Yes | Yes | Yes | (Yes) | Yes | Museums* (5); pubs; eDNA | |
| Belonidae | uncertain | Tylosurus fodiator# | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; eDNA | |
| Belonidae | yes | Tylosurus melanotus# | Yes | Yes | Yes | Yes | Yes | Museums* (2); eDNA | |
| Blenniidae | yes | Entomacrodus chiostictus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); inat; pubs; 2022- images, specimens |
| Blenniidae | yes | Hypsoblennius proteus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); pubs; inat; 2022- images, specimens |
| Blenniidae | yes | Ophioblennius clippertonensis# | Yes | Yes | Yes | Yes | Museum (1); inat?; DNA; 2022- images, specimens, DNA | ||
| Blenniidae | yes | Ophioblennius steindachneri# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); inat?; pubs; 2022- images, specimens |
| Blenniidae | yes | Plagiotremus azaleus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (3); observe; 2022- images |
| Bothidae | yes | Bothus leopardinus | Yes | Yes | Yes | Yes | Yes | Museums (3); pubs; inat; 2022- images, specimen | |
| Bothidae | yes | Bothus mancus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); observe; pubs; inat; 2022- images, specimen |
| Bythitidae | yes | Grammonus diagrammus# | Yes | Yes | Yes | Museums* (2) | |||
| Carangidae | no | Alectis ciliaris# | Yes | Yes | Yes | Museum* (1); images | |||
| Carangidae | uncertain | Caranx caballus# | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums (4); observe; inat singleton, small group; 2022- images, singletons |
| Carangidae | no | Caranx caninus# | Yes | Yes | (Yes) | Yes | observe | ||
| Carangidae | yes | Caranx lugubris | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); inat; observe; pubs; 2022- images, common |
| Carangidae | yes | Caranx melampygus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); observe; inat; pubs; 2022- images, common |
| Carangidae | yes | Caranx sexfasciatus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); inat; observe; pubs; 2022- images, schools |
| Carangidae | yes | Decapterus macarellus | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums (4); inat; 2022- images, large schools |
| Carangidae | yes | Decapterus muroadsi | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums* (5); pub |
| Carangidae | yes | Elagatis bipinnulata | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; pub; observe; 2022- images, common |
| Carangidae | yes | Ferdauia orthogrammus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); inat; observe; pub; 2022- images, small schools |
| Carangidae | yes | Naucrates ductor | Yes | Yes | Yes | Yes | (Yes) | Yes | Museum (1); observe; inat |
| Carangidae | yes | Selar crumenophthalmus | Yes | Yes | (Yes) | Yes | Museums* (4); pubs | ||
| Carangidae | no | Selene peruviana# | Yes | (Yes) | Museums* (2) | ||||
| Carangidae | no | Seriola dorsalis# | Yes | Yes | Yes | Yes | Yes | Four: observe; eDNA | |
| Carangidae | yes | Seriola rivoliana | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums (5); pubs; inat; 2022- observe, image |
| Carangidae | yes | Trachinotus stilbe | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); pubs; inat; observe; 2022- images, schools |
| Carangidae | no | Trachurus symmetricus | Yes | (Yes) | Yes | Pub | |||
| Carangidae | yes | Uraspis helvola | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums (3); inat; 2022- images, schools |
| Carcharhinidae | no | Carcharhinus altimus | Yes | (Yes) | Yes | Museum*(1) | |||
| Carcharhinidae | yes | Carcharhinus albimarginatus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (3); observe; inat; pubs; 2022- images, common |
| Carcharhinidae | yes | Carcharhinus falciformis | Yes | Yes | Yes | Yes | (Yes) | Yes | Museum* (1); pubs; inat; observe; 2022- images, common |
| Carcharhinidae | yes | Carcharhinus galapagensis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); observe; inat; pubs; 2022- images, common |
| Carcharhinidae | yes | Carcharhinus limbatus | Yes | Yes | Yes | Yes | Yes | Museums (3); observe; pubs; inat | |
| Carcharhinidae | uncertain | Carcharhinus obscurus | Yes | Yes | Yes | Yes | Pubs; capture; observe | ||
| Carcharhinidae | no | Nasolamia velox | Yes | Yes | Yes | (Yes) | Yes | Pub; capture | |
| Carcharhinidae | yes | Triaenodon obesus | Yes | Yes | Yes | Yes | Yes | Yes | Pubs; observe; inat; 2022- Images, common: |
| Chaenopsidae | yes | Acanthemblemaria mangognatha | Yes | Yes | Yes | Yes | Yes | Museums* (6); pub; 2022- images, specimens, common | |
| Chaenopsidae | uncertain | Chaenopsis alepidota# | Yes | 2022- image | |||||
| Chaetodontidae | no | Chaetodon humeralis# | Yes | Yes | Yes | Yes | Museums (2); pub; inat | ||
| Chaetodontidae | yes | Forcipiger flavissimus# | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (8); pubs; inat; observe; 2022- images, specimens |
| Chaetodontidae | yes | Johnrandallia nigrirostris | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); observe; inat; pubs; 2022- images, common |
| Chaetodontidae | yes | Prognathodes falcifer# | Yes | Yes | Yes | Yes | Yes | Yes | Pub; observe; images, common |
| Cirrhitidae | yes | Cirrhitichthys oxycephalus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); pub; observe; inat; 2022- images, specimens, common |
| Cirrhitidae | yes | Cirrhitus rivulatus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (8); inat; pubs; 2022- images, specimens, common |
| Congridae | yes | Paraconger californiensis | Yes | Yes | Yes | Yes | Yes | Museums* (2) | |
| Congridae | uncertain | Paraconger similis | Yes | Yes | Yes | Museum* (1) | |||
| Cyclopsettidae | uncertain | Citharichthys sp. | Yes | Image | |||||
| Cyclopsettidae | no | Syacium ovale | Yes | (Yes) | Yes | Museum* (1) | |||
| Cynoglossidae | uncertain | Symphurus atramentatus | Yes | (Yes) | Yes | Museum* (1) | |||
| Dactyloscopidae | yes | Dactyloscopus insulatus | Yes | Yes | Yes | Yes | Yes | Museums (4); pub; 2022- image, specimen, DNA | |
| Dactyloscopidae | yes | Gillellus semicinctus | Yes | Yes | Yes | Yes | Museums (2); image; 2022- specimens, DNA | ||
| Dactyloscopidae | yes | Myxodagnus opercularis# | Yes | Yes | Yes | Yes | Museums* (3); 2022- specimen, DNA | ||
| Dasyatidae | no | Hypanus dipterurus# | Yes | Yes | Yes | Yes | Museum (1); pub; image | ||
| Dasyatidae | yes | Hypanus longus# | Yes | Yes | Yes | Yes | Yes | Museums (2); pub; inat; observe; 2022- images | |
| Diodontidae | yes | Chilomycterus reticulatus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; observe; 2022- images |
| Diodontidae | yes | Diodon holocanthus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); observe; inat; pub; 2022- images, specimens |
| Diodontidae | yes | Diodon hystrix | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); pubs; observe; 2022- images |
| Echeneidae | uncertain | Echeneis naucrates | Yes | Yes | Yes | Museum | |||
| Echeneidae | yes | Phtheirichthys lineatus | Yes | (Yes) | Yes | Museums* (3) | |||
| Echeneidae | yes | Remora remora | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums (4); inat; pubs; observe; 2022- images |
| Ephippidae | no | Platax teira | Yes | Yes | Images: 2025(2, 1 from inat) | ||||
| Epinephelidae | yes | Alphestes immaculatus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; observe; 2022- images, specimens |
| Epinephelidae | yes | Cephalopholis colonus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); inat; observe; pubs; 2022- images, specimens, common |
| Epinephelidae | yes | Cephalopholis panamensis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); inat; observe; pubs; 2022- images, specimens, common |
| Epinephelidae | yes | Dermatolepis dermatolepis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); pubs; inat; 2022- images, specimen |
| Epinephelidae | no | Epinephelus analogus# | Yes | Yes | Yes | Museum* (1); pubs | |||
| Epinephelidae | yes | Epinephelus clippertonensis# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; DNA; 2022- images, common |
| Epinephelidae | yes | Epinephelus labriformis# | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (5); inat; DNA; 2022- images, uncommon |
| Epinephelidae | yes | Hyporthodus cifuentesi# | Yes | Yes | Yes | Yes | Yes | Yes | Museum (1); pubs; images |
| Fistulariidae | yes | Fistularia commersonii | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (3); pubs; inat; observe; 2022- images |
| Fistulariidae | uncertain | Fistularia corneta | Yes | Yes | Yes | Yes | Yes | Museums* (2) | |
| Galeocerdonidae | yes | Galeocerdo cuvier | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); pubs; observe; inat; 2022- images |
| Gobiesocidae | yes | Gobiesox canidens# | Yes | Yes | Yes | Yes | Yes | Museums (3); pubs; DNA; inat; 2022- images, specimens, DNA | |
| Gobiesocidae | yes | Tomicodon absitus# | Yes | Yes | Yes | Yes | Yes | Museums (3); pubs; inat; DNA; 2022- images, specimens, DNA | |
| Gobiesocidae | yes | Tomicodon sp.# | Yes | Museum (1); DNA; 2022- images, specimens | |||||
| Gobiesocidae | no | Tomicodon petersii | Yes | Yes | Museum* (1); pub | ||||
| Gobiidae | yes | Bathygobius ramosus longipinnis# | Yes | Yes | Yes | Yes | Museums* (5); pubs; inat; DNA | ||
| Gobiidae | yes | Chriolepis sp.# | Yes | Yes | Museums (3); 2022- images, specimens, DNA | ||||
| Gobiidae | yes | Coryphopterus urospilus# | Yes | Yes | Yes | Yes | Yes | Museums (3); pub; DNA; 2022- images, specimens, DNA | |
| Gobiidae | yes | Lythrypnus cf. dalli# | Yes | 2022- images, specimens, DNA | |||||
| Gobiidae | yes | Lythrypnus insularis# | Yes | Yes | Yes | Yes | Yes | Museums (5); pub; 2022- images, specimens, DNA | |
| Gobiidae | yes | Schindleria praematura | Yes | Yes | Yes | Yes | Museums (5); eDNA 2022- specimen, DNA | ||
| Grammistidae | yes | Pseudogramma thaumasia | Yes | Yes | Yes | Yes | Yes | Yes | Museums (3); 2022- image, specimens |
| Grammistidae | yes | Rypticus courtenayi# | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (7); DNA; inat; pubs; 2022- images, specimens, DNA |
| Grammistidae | no | Rypticus nigripinnis# | Yes | Yes | Yes | Museum* (1) | |||
| Haemulidae | yes | Anisotremus perezponcedeleoni# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); DNA; pubs; inat; 2022- images, specimens |
| Haemulidae | no | Anisotremus taeniatus# | Yes | Yes | Yes | Observe | |||
| Haemulidae | no | Orthopristis cantharina# | Yes | Image | |||||
| Hemiramphidae | uncertain | Hemiramphus saltator# | Yes | (Yes) | Yes | Inat | |||
| Holocentridae | yes | Myripristis berndti | Yes | Yes | Yes | Yes | Yes | Yes | Museum (1); inat; 2022- images, specimens |
| Holocentridae | yes | Myripristis clarionensis# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (9); inat; pubs; 2022- images, specimens, common |
| Holocentridae | no | Myripristis leiognathus# | Yes | Yes | Yes | Yes | Museum* (1); pub; 2022- specimen, image | ||
| Holocentridae | yes | Neoniphon suborbitalis | Yes | Yes | Yes | Yes | Yes | Yes | Museum (5); pubs; 2022- images, specimens |
| Holocentridae | yes | Plectrypops lima | Yes | Yes | Yes | Yes | Museum (1); 2022- image, specimen | ||
| Kuhliidae | yes | Kuhlia mugil | Yes | Yes | Yes | Yes | Yes | Museums (8); pubs; inat; 2022- images, specimens, DNA, schools | |
| Kyphosidae | yes | Kyphosus elegans | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); pub, observe; inat; 2022- images, specimens, common |
| Kyphosidae | yes | Kyphosus ocyurus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); observe; 2022- images, common |
| Kyphosidae | yes | Kyphosus sectatrix# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); pubs; observe; inat; 2022- images, specimens, common |
| Kyphosidae | yes | Kyphosus vaigiensis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); observe; inat; pub; 2022- image. specimens, observe |
| Labridae | yes | Bodianus diplotaenia | Yes | Yes | Yes | Yes | Yes | Yes | Museums (8); observe; pubs; inat; 2022- images, specimens, common |
| Labridae | uncertain | Halichoeres adustus# | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; pubs; 2022- image, rare | |
| Labridae | yes | Halichoeres insularis# | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (3); inat; observe; pubs; 2022- images, specimens, DNA |
| Labridae | yes | Halichoeres nicholsi# | Yes | Yes | Yes | Yes | Yes | Museums (7); pubs; inat; DNA; 2022- images, specimens, common | |
| Labridae | yes | Halichoeres notospilus | Yes | Yes | Yes | Yes | Museums (5); pubs; inat; 2022- images, observe | ||
| Labridae | yes | Halichoeres sanchezi# | Yes | Yes | 2022- images, specimens | ||||
| Labridae | uncertain | Iniistius pavo | Yes | Yes | Yes | Yes | Yes | Observe; 2022- images | |
| Labridae | yes | Novaculichthys taeniourus | Yes | Yes | Yes | Yes | Yes | Observe; inat; 2022- images, specimens | |
| Labridae | yes | Stethojulis bandanensis# | Yes | Yes | Yes | Yes | Yes | Museum (1); inat; eDNA; 2022- images, observe | |
| Labridae | yes | Thalassoma grammaticum# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (8); pubs; observe, inat; 2022- images, specimens, common |
| Labridae | yes | Thalassoma lucasanum | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); inat; observe; pubs; 2022- images, common |
| Labridae | no | Thalassoma purpureum | Yes | 2022- image, single individual | |||||
| Labridae | yes | Thalassoma virens# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); inat; pubs; observe; 2022- images |
| Labridae | yes | Xyrichtys sp.# | Yes | Yes | Yes | Yes | Images; observe, pub | ||
| (Scarinae) | yes | Calotomus carolinus | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (6); inat; pubs; observe; 2022- images, specimens, common |
| (Scarinae) | no | Scarus compressus | Yes | Yes | Yes | Museum (3); pub; 2022- image | |||
| (Scarinae) | no | Scarus ghobban | Yes | Yes | Yes | Yes | Museum (1); pub; inat; observe; 2022- image | ||
| (Scarinae) | yes | Scarus rubroviolaceus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); observe; pub; inat; 2022- images, specimens, DNA, common |
| Labrisomidae | yes | Labrisomus socorroensis# | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (5); pubs; inat; 2022- image |
| Labrisomidae | no | Malacoctenus hubbsi | Yes | Museum* (1) | |||||
| Lamnidae | no | Carcharodon carcharias | Yes | Yes | Yes | Pub | |||
| Latilidae | uncertain | Caulolatilus affinis | Yes | Yes | Yes | Yes | Museum* (1) | ||
| Latilidae | uncertain | Caulolatilus princeps | Yes | Yes | Yes | Pub | |||
| Lobotidae | no | Lobotes pacifica | Yes | (Yes) | Yes | Museum* (1) | |||
| Lutjanidae | no | Hoplopagrus guentherii# | Yes | Yes | Yes | Pub | |||
| Lutjanidae | no | Lutjanus aratus | Yes | 2022- specimen; DNA | |||||
| Lutjanidae | no | Lutjanus argentiventris# | Yes | Yes | Yes | Yes | Yes | Pub; observe; image; 2022- image | |
| Lutjanidae | no | Lutjanus inermis# | Yes | Yes | Yes | Pub | |||
| Lutjanidae | yes | Lutjanus peru | Yes | Yes | Yes | Museum* (1); pub; image | |||
| Lutjanidae | yes | Lutjanus viridis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (9); pubs; inat; 2022- images, specimens, common |
| Mobulidae | yes | Mobula birostris | Yes | Yes | Yes | Yes | (Yes) | Yes | Museum (1); pubs, inat, observe; 2022- images, common |
| Mobulidae | no | Mobula tarapacana# | Yes | Yes | Yes | (Yes) | Yes | Pub; observe; inat | |
| Mobulidae | no | Mobula thurstoni# | Yes | Yes | Yes | Image; eDNA; 2022- image | |||
| Monacanthidae | uncertain | Aluterus monoceros# | Yes | Yes | Yes | Yes | Yes | Museums (1); pub; observe; inat; 2022- images | |
| Monacanthidae | yes | Aluterus scriptus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); pubs; inat; observe; 2022- images |
| Monacanthidae | yes | Cantherhines dumerilii | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); pubs; observe; inat; 2022- images, specimens |
| Mugilidae | yes | Chaenomugil proboscideus | Yes | Yes | Yes | Yes | Yes | Museums (9); pubs; inat; 2022- specimens, DNA | |
| Mugilidae | uncertain | Mugil cephalus# | Yes | Inat | |||||
| Mugilidae | yes | Mugil setosus# | Yes | Yes | Yes | Yes | Museums* (6), Image; pubs | ||
| Mullidae | yes | Mulloidichthys dentatus | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (7); pubs; inat; observe; 2022- images, specimens |
| Muraenidae | yes | Anarchias galapagensis | Yes | Yes | Yes | Yes | Museums* (3) | ||
| Muraenidae | yes | Echidna nebulosa# | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; 2022- images | |
| Muraenidae | yes | Echidna nocturna# | Yes | Yes | Yes | Yes | Yes | Museums (2); pubs; inat; 2022- observe | |
| Muraenidae | yes | Enchelycore octaviana | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; observe; 2022- image, observe |
| Muraenidae | yes | Gymnomuraena zebra | Yes | Yes | Yes | Yes | Yes | Museum (1); pub; inat; observe; 2022- image | |
| Muraenidae | yes | Gymnothorax castaneus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (3); inat; pub; observe; 2022- images |
| Muraenidae | yes | Gymnothorax dovii | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); pub; observe; inat; 2022- images |
| Muraenidae | no | Gymnothorax flavimarginatus# | Yes | Yes | Yes | Yes | Museum (1); eDNA; 2022- image | ||
| Muraenidae | yes | Gymnothorax panamensis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); inat; pub; 2022- image, specimens, DNA |
| Muraenidae | no | Gymnothorax pictus | Yes | Yes | Yes | Museum* (1); pubs | |||
| Muraenidae | no | Gymnothorax verrilli | Yes | 2022- specimen, DNA | |||||
| Muraenidae | yes | Muraena lentiginosa | Yes | Yes | Yes | Yes | Yes | Museums (3); image; pubs; observe | |
| Muraenidae | yes | Scuticaria tigrina | Yes | Yes | Yes | Yes | Yes | Yes | Museums (3); pub; eDNA; 2022- images |
| Muraenidae | yes | Uropterygius macrocephalus | Yes | Yes | Yes | Yes | Museums* (5); Image; inat; pub; 2022- specimen, DNA | ||
| Narcinidae | yes | Diplobatis ommata | Yes | Yes | Pub; inat; 2022- observe | ||||
| Narcinidae | yes | Narcine entemedor | Yes | Yes | Yes | Yes | Yes | Museums (3); pubs; inat; 2022- images, observe | |
| Ophichthidae | yes | Ichthyapus selachops | Yes | Yes | Yes | Museum* (1); DNA | |||
| Ophichthidae | yes | Myrichthys pantostigmius# | Yes | Yes | Yes | Yes | Yes | Museum* (3); inat; pubs; 2022- images | |
| Ophichthidae | uncertain | Paraletharchus opercularis | Yes | Yes | Yes | Museum* (1) | |||
| Ophichthidae | yes | Quassiremus evionthas | Yes | Pub; images; 2022- image | |||||
| Ophidiidae | yes | Brotula ordwayi# | Yes | Yes | Yes | Yes | Museums* (2); image | ||
| Ostraciidae | yes | Lactoria diaphana | Yes | Yes | Yes | Yes | Yes | Museums* (3); inat; 2022- image, specimen | |
| Ostraciidae | yes | Ostracion meleagris | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); inat; pubs; 2022- images, specimens |
| Pomacanthidae | yes | Holacanthus clarionensis# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (9); inat; pubs; 2022- images, specimens, common |
| Pomacanthidae | no | Holacanthus passer# | Yes | Yes | Yes | Yes | Yes | Yes | Museum (2); image |
| Pomacanthidae | no | Pomacanthus zonipectus | Yes | Yes | Yes | Museum* (1); observe | |||
| Pomacentridae | yes | Abudefduf troschelii | Yes | Yes | Yes | Yes | Yes | Museums (9); pubs; inat; 2022- images, specimens, common | |
| Pomacentridae | uncertain | Azurina atrilobata# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (2); pubs, observe; 2022- image, specimen, DNA |
| Pomacentridae | yes | Azurina hirundo# | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; pubs; observe; image | |
| Pomacentridae | yes | Chromis alta | Yes | Yes | Yes | Yes | Yes | Yes | Museum (1); pub; images; 2022- image |
| Pomacentridae | yes | Microspathodon bairdii | Yes | Yes | Yes | Yes | Yes | Museums (5); Images; pubs; inat; 2022- images | |
| Pomacentridae | yes | Microspathodon dorsalis | Yes | Yes | Yes | Yes | Yes | Yes | Museums (9); inat; pubs; observe; 2022- images, specimens |
| Pomacentridae | no | Stegastes acapulcoensis# | Yes | Yes | Yes | 2022- image, specimens, DNA | |||
| Pomacentridae | no | Stegastes flavilatus# | Yes | Yes | Yes | Yes | Yes | Yes | Pub; inat; observe; 2022- image, specimens, DNA |
| Pomacentridae | yes | Stegastes leucorus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (7); pubs; inat; observe; 2022- images, specimens, common |
| Pomacentridae | no | Stegastes rectifraenum | Yes | Yes | Yes | Yes | Museums (1): mid-20th century only | ||
| Pomacentridae | yes | Stegastes redemptus# | Yes | Yes | Yes | Yes | Yes | Museums (8); inat; pubs; observe; 2022- images, specimens, DNA, common | |
| Priacanthidae | yes | Cookeolus japonicus | Yes | Yes | Yes | Yes | Yes | Museums (3), pub | |
| Priacanthidae | yes | Heteropriacanthus carolinus# | Yes | Yes | Yes | Yes | Yes | Museums* (5); pubs; inat; 2022- images, specimens | |
| Priacanthidae | yes | Priacanthus alalaua# | Yes | Yes | Yes | Yes | Yes | Yes | Museum* (1); pubs; image |
| Priacanthidae | yes | Pristigenys serrula# | Yes | Yes | Yes | Pubs; image | |||
| Rhincodontidae | yes | Rhincodon typus | Yes | Yes | Yes | Yes | (Yes) | Yes | Pub; inat; observe 2022- images |
| Sciaenidae | yes | Pareques sp.# | Yes | Yes | Yes | Yes | Yes | Museums (4); pub; DNA; 2022- images, specimens, DNA | |
| Scombridae | yes | Acanthocybium solandri | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums (2); inat; observe; eDNA 2022- images, observe |
| Scombridae | yes | Auxis brachydorax | Yes | Yes | Yes | (Yes) | Yes | Museum (1); eDNA; 2022- image, school | |
| Scombridae | no | Euthynnus affinis | Yes | (Yes) | Yes | Museum* (1); observe | |||
| Scombridae | yes | Euthynnus lineatus | Yes | Yes | Yes | Yes | (Yes) | Yes | Museums (3); inat; observe; 2022- image, school, observe |
| Scombridae | yes | Katsuwonus pelamis | Yes | Yes | (Yes) | Yes | Museum (1); pubs; eDNA; 2022- image, school | ||
| Scombridae | yes | Scomber australasicus# | Yes | Yes | Yes | (Yes) | Yes | Museum* (1); pubs | |
| Scombridae | no | Scomber japonicus# | Yes | Yes | (Yes) | Yes | Museum (1); pubs | ||
| Scombridae | Yes | Thunnus albacares | Yes | Yes | Yes | Yes | (Yes) | Yes | Museum (1); observe; pubs; inat; 2022- images; observe, DNA |
| Scorpaenidae | yes | Pontinus vaughani | Yes | Yes | Yes | Yes | Yes | Yes | Museum* (2); images |
| Scorpaenidae | yes | Scorpaena afuerae# | Yes | Yes | Images | ||||
| Scorpaenidae | uncertain | Scorpaena histrio | Yes | Yes | Yes | Yes | Pub; image | ||
| Scorpaenidae | yes | Scorpaena mystes | Yes | Yes | Yes | Yes | Yes | Museums* (4); inat; pubs; 2022- images | |
| Scorpaenidae | yes | Scorpaenodes xyris# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (3); pubs; 2022- images, specimens |
| Serranidae | uncertain | Serranus aequidens# | Yes | Yes | Yes | Museum (1); eDNA | |||
| Serranidae | yes | Serranus socorroensis# | Yes | Yes | Yes | Yes | Yes | Yes | Museums* (3); pubs; observe; 2022- images, specimens |
| Soleidae | yes | Aseraggodes herrei# | Yes | Yes | Yes | (Yes) | Yes | Museums (3); Image; 2022- images, specimens, DNA | |
| Sphyrnidae | yes | Sphyrna lewini | Yes | Yes | Yes | Yes | Yes | Yes | Museum (1); pubs; inat; observe; 2022- images, observe |
| Syngnathidae | yes | Bryx veleronis# | Yes | Yes | Yes | Yes | Museums* (2); pubs; 2022- specimens, DNA | ||
| Syngnathidae | yes | Doryrhamphus melanopleura# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (4); pubs; DNA; 2022- images, specimens, observe, DNA |
| Syngnathidae | no | Syngnathus auliscus | Yes | (Yes) | Yes | Museum* (1) | |||
| Synodontidae | uncertain | Synodus lacertinus | Yes | Yes | Yes | Yes | Yes | Museum (1); 2022- images, specimens, DNA | |
| Tetraodontidae | yes | Arothron hispidus | Yes | Yes | Yes | Yes | Yes | Museum (1); observe; 2022- images, specimens | |
| Tetraodontidae | yes | Arothron meleagris | Yes | Yes | Yes | Yes | Yes | Yes | Museums (10); inat; pubs; 2022- images, specimens, common |
| Tetraodontidae | yes | Canthigaster janthinoptera# | Yes | Yes | Museum (1); inat; DNA; 2022- images | ||||
| Tetraodontidae | yes | Canthigaster punctatissima# | Yes | Yes | Yes | Yes | Yes | Yes | Museums (5); inat; pubs; DNA; 2022- images, specimens, DNA |
| Tetraodontidae | yes | Sphoeroides lobatus | Yes | Yes | Yes | Yes | Yes | Museums (2); inat; pub; 2022- images, specimens | |
| Triglidae | no | Bellator loxias | Yes | (Yes) | Yes | Pub | |||
| Tripterygiidae | yes | Axoclinus multicinctus | Yes | Yes | Yes | Yes | Yes | Museums (3); inat; pub; 2022- images, specimens; DNA | |
| Tripterygiidae | yes | Enneanectes exsul | Yes | Yes | Yes | Yes | Yes | Museums (4); pub; 2022- images, specimens, DNA | |
| Zanclidae | yes | Zanclus cornutus | Yes | Yes | Yes | Yes | Yes | Yes | Museums (6); observe; inat; pubs; 2022- images, specimens, common |
DNA sequencing of specimens, single specimens per species in most cases, collected during the 2022 expedition provided confirmation of the identity of individuals of 35 species. More expansive genetic studies by ODD’s laboratory, some of them previously published (Acevedo-Álvarez et al., 2021; Bernal-Hernández et al., 2024; Torres-García et al., 2024; Torres-Hernández et al., 2021, 2022) and some in progress, provided information on the population status of 15 species in relation to endemism. Both those sets of genetic data provided useful information about the endemism status of RNP species. UNESCO eDNA provided supporting data for 2 old single museum collections (Pronotogramus multifasciatus, the only member of its genus in the TEP, and Serranus aequidens) and 2 species that only had observational records (Ablennes hians and Seriola dorsalis). In addition, 2023 eDNA data provided additional information on the distributions of 11 other species among the different RNP islands that were already confirmed by other primary data as present. Finally, we accepted observation-only records of 5 species with distinctive morphology (Carcharodon carcharhias, Lutjanus inermis, Anisotremus taeniatus, Hoplopagrus guentheri, and Caranx caninus). Based on all sources of data there are 189 species known from Clarion, 126 from Roca Partida, 159 from San Benedicto and 204 from Socorro (Tables 1, Supplementary material: Table S1).
Photographic database of accepted species
The 9 photographers on the 2022 expedition produced more than 5,000 underwater images of living fishes in their natural habitats. Those images confirmed the occurrences of 137 species at various RNP islands, with another 9 species confirmed by specimen images. Similar images from 2 other coauthors (CDC and AAB), taken at Roca Partida, San Benedicto and Socorro in 2015, confirmed 25 island occurrences (Supplementary material: Files S2-4). Live-fish images taken over the past decade by 19 iNaturalist underwater photographers also provided confirmation of 123 species occurrences, among them vouchers of 13 species not photographed during the 2022 expedition, with 71 of those images being included in the 4 island-fauna supplemental plates (Supplementary material: Files S1-S4). Those iNaturalist images provided sole-source documentation of occurrence of 3 species in the RNP, Mobula tarapacana (previously recorded as observed), Hemiramphus saltator and Mugil cephalus. The deep-reef videos of the ROV, BRUV and submersible studies provided voucher images in the supplemental island-fauna plates for 23 occurrences at different islands of 16 species and were the only sources of images for 11 of those species (Aburto-Oropeza et al., 2016; Ayala-Bocos et al., 2015; Hollarsmith et al., 2020). Altogether those various sources provided images of 181 species in 125 genera and 57 families, including 149 species of residents, 144 of them with live-fish images. Images provided the only voucher record of 5 species (Chaenopsis alepidota, Citharichthys sp., Platax teira, Scorpaena afuerae, and Thalassoma purpureum) newly added to the fauna. The 4 island-fauna image plates in Supplementary material: files S1-S4 include 468 field images of living fishes plus 14 of specimens collected there in 2022. Together all those images provide confirmation of 123 species from Clarion (Supplementary material: File S1), 74 from Roca Partida (Supplementary material: File S2), 118 from San Benedicto (Supplementary material: File S3) and 151 from Socorro (Supplementary material: File S4).
Species accounts of accepted species
These accounts expand on the information contained on those species in the 2 “accepted species” tables (Tables 1, Supplementary material: ST1), providing details relating to the occurrence and population status of various species of reef-associated fishes at different islands in the RNP.
Ablennes hians. This pantropical, epipelagic needlefish occurs primarily in offshore areas and is often found inshore around islands (Collette et al., 2015). It is the only member of its genus. In the TEP it is distributed from the Gulf of California and southern Baja to northern Peru, plus the Galápagos, Cocos and Malpelo. The 2002 checklist and both 2016 inventories included A. hians, but without providing primary source information. Although there are relatively few museum records and iNaturalist images of this species in the TEP, that likely reflects the species use of offshore habitats rather than rarity. While there is only one museum record from the Galápagos, Grove and Lavenberg (1997) recorded large schools occurring seasonally in inshore waters of that archipelago. The only museum record relating to the RNP is an SIO specimen of Tylosurus fodiator, misidentified as A. hians, collected just outside the park limits. Coauthor AAB observed small schools (< 10 individuals) of A. hians at a single, high wave-energy site at San Benedicto in 2017, 2019, and 2022. The eDNA expedition of 2023 produced 603 reads of 1 ASV of CO1 plus tens of thousands of reads of multiple ASVs of 16S (100% species-name match in both cases) from Clarion, San Benedicto, and Socorro. Given these observations and the amounts of DNA recovered at the 3 main islands we class A. hians as a resident.
Acanthurus xanthopterus. This Indo-Pacific species is widespread in the TEP, from the Gulf of California to Ecuador and all the oceanic islands. All records of this species in the RNP are recent, from 2007 (Chávez- Comparan et al., 2010) to 2024, with the most multiyear records represented by observations by AAB at Roca Partida, plus the collection of a single specimen by ODD. All images of this readily recognizable, large (to 70 cm TL) species that are available, from both the 2022 expedition and iNaturalist, are of solitary, medium-large to large adults (Supplementary material: Files S1-S4), except for 1 of 2 adults by iNaturalist, and most are from Roca Partida. This species is known to live for up to at least 35 years (Taylor et al., 2024) and what might have been the same individual, was observed in midwater over multiple years at that tiny speck of habitat, likely feeding on the feces of carnivorous and planktivorous fishes (Abesamis et al., 2012), which are particularly abundant at Roca Partida. We tentatively class it as a resident.
Acanthemblemaria mangognatha. See endemics section A.
Alectis ciliaris. The only museum record of this large benthopelagic species is of a single individual collected in 1980 (LACM; specimen ID confirmed) at Socorro. AAB obtained images of 2 adults at San Benedicto in 2023 and observed single adults at that same island in 2017 and 2019 and at Socorro in 2023 and there is an iNaturalist image from Socorro (Supplementary material: File S4). These scattered observations of so few adults and no juveniles of a species known to often form schools do not indicate the presence of a resident population.
Aluterus monoceros. This circumtropical monacanthid is widely distributed in the TEP, from the Gulf of California to Chile and all the offshore TEP islands except Clipperton. The only RNP museum records are of 4 fish collected at San Benedicto in 1951 that are in SIO. Single individuals were photographed at San Benedicto and Socorro in 2022 and AAB observed single individuals at San Benedicto in 2013, Roca Partida in 2014 and Socorro in 2017 and 2023. Fourriére et al. (2016) observed it in 2012-2013 but not in 1994-1999, 2007 or 2010. These records scattered across a large range of years refer to few individuals. This species also associates with offshore flotsam (Hunter & Mitchell, 1966, 1968) and has been videoed at near- surface BRUVS deployed in deep water between Cocos and the Galápagos (Cambra et al., 2021; Vaudo et al., 2023; and see Amonoceros_floatinglog). This pelagic habit may lead to occasional recruitment to the RNP. Due to these issues, we record its population status as uncertain.
Anarchias galapagensis. This small, highly cryptic moray eel is rarely seen or collected without the use of ichthyocides. It has been collected at Socorro in 1939 (2 individuals LACM), 1971 (3 individuals LACM) and 1995 (1 individual CNPE) and at Clarion in 1959 (1 individual LACM), 1971 (10 individuals LACM) and 1974 (1 individual SIO). Since it was collected as recently as 1995 and there were multiple collections at 2 islands over several decades we class this species as a resident.
Anisotremus perezponcedeleoni. See endemics section A.
Anisotremus taeniatus. Both adults and juveniles of A.taeniatus, a common and widely distributed species in the TEP, have very distinctive, conspicuous color patterns that make them readily distinguishable from all other similar species. Hence, although the only record of this species is from an observation, we accept it as a validated occurrence in the RNP, as a vagrant.
Aseraggodes herrei. See endemics section C.
Axoclinus multicinctus. See endemics section A.
Azurina atrilobata. This easily recognized species is abundant on shallow reefs on the mainland, Cocos and the Galápagos and has been recorded only a few times in the RNP. There are specimens at CMNFI collected at Socorro in 1957, and a single specimen at LACM caught in deep water ~ 4 km offshore from San Benedicto (in Baja State in the database, rather than Colima State where the RNP is located), AAB observed it at Roca Partida and San Benedicto in 2012, but not 2013, 2014, 2017, 2019, 2020 or 2023. Chávez-Comparan et al. (2010) did not include this species among those they recorded at Socorro in 2007. It was photographed and a specimen collected from a small group of adults at San Benedicto in 2022 and another collected at Clarion (DNA barcoded) but not encountered at the other 2 islands. This species has been found in abundance associated with offshore flotsam (Hunter & Mitchell, 1966, 1968; Mora et al., 2001) which could facilitate its arrival in groups in the RNP. Given the paucity and intermittency of records it is regarded as an uncertain resident in the RNP.
Azurina hirundo. This species, which was described from specimens collected at Guadalupe Island, is distributed from southern California (where it now appears to have a resident population) to islands scattered along the coast of Baja California (Guadalupe, the Cedros island complex, and Alijos Rocks) and the RNP. Adults of this distinctive, readily recognizable species have a long history of being observed in the 3 easternmost islands, between the early 1990s to the early 2010s (Fourriére et al., 2016), as well as 2016, 2017, 2023, and 2024. Photographs of this species in the RNP are available from 1994, 2015, 2023, and 2024, the last 2 of groups of 9-10 individuals. However, it was not observed at any of the 4 islands during the 2022 expedition. Although its known depth range is fairly shallow, 5~30 m, this active, midwater feeding planktivore may sometimes range into deeper water when feeding and hence was missed by scuba divers in 2022. Azurina atrilobata, for example, is found as deep as 80 m, and Azurina multilineata from the Caribbean down to 130 m. However, Hollarsmith et al. (2020) used an ROV to survey 15 sites between 19-80 m depth at San Benediction and Socorro in 2018 but recorded no A. hirundo. In June 2024, AAB observed thousands of juveniles of this species in shallow water at Roca Partida, which, when he returned in November, had reached about 5 cm TL. We conclude that this species, which we class as a resident, undergoes substantial population fluctuations in the RNP. However, it is also possible that the RNP population is maintained to some extent by influxes of larvae carried south from more northerly populations by the California Current.
Bathygobius ramosus longipinnis. See endemics section B.
Brotula ordwayi. This large ophidiid (reaches 75 cm TL) has been collected once at Roca Partida (LACM, 1939, 1 adult) and an individual was videoed at Socorro in 2015 by CAS (Supplementary material: File S4). There are no other records. This is a deepwater species typically found well below scuba depth limits (depth range to 270 m)and neither Hollarsmith et al. (2020) nor Giddens et al. (2019) recorded this species in their video studies of fishes in deep water in the RNP. Hence, we record its population status as uncertain.
Bryx clarionensis. See endemics section D.
Canthidermis maculata. This semioceanic species is known from the mouth of the Gulf of California to Colombia, plus all the oceanic islands, and from oceanic records, where it associates with flotsam (Hunter & Mitchell, 1966) and BRUVS (Vaudo et al., 2023). There are 2 SIO specimens, collected in 1962 (several kilometers offshore from Socorro) and 1975, plus 1 more recently that is in CPUM. Berry and Baldwin (1966) also recorded it from the RNP. The occasional collection of single individuals over 60 + years indicates a persistent presence, but with an uncertain residency status due to its pelagic habits and scarcity.
Canthigaster punctatissima and Canthigaster janthinoptera. According to Allen and Randall (1977), C. punctatissima has the following species-specific coloration: a brown head and body covered in numerous small, round white to blue-white spots that extend onto the base of the caudal fin, no short pale lines radiating from the eye and no entire or partial ocellus at the base of the dorsal fin. According to those authors the only morphological and morphometric difference between C. punctatissima and the Indo-Central Pacific (ICP) species, C.janthinoptera is their coloration. In addition to a body covered in small round spots like those of C. punctatissima that can also extend onto the base of the caudal fin, C. janthinoptera has pronounced pale lines radiating from the eyes and a large, entire or partial ocellus at the base of the dorsal fin (Allen & Erdmann, 2024; Allen & Randall, 1977). Based on images of C. janthinoptera from the ICP on iNaturalist this species varies geographically in coloration in that large region, with the addition of thin pale lines on other parts of the head and parts of the body in some populations. There are numerous images of Canthigaster at various sites in the TEP that fit the definition of C. janthinoptera. At the RNP small, recently recruited juveniles, whose coloration was not mentioned by either Allen and Erdmann (2024) or Allen and Randall (1977), have a well-defined ocellus at the base of the dorsal fin, lines on the head and longitudinally on the nape and many fewer, more dispersed round spots on the sides of the body than is seen in adults of both species.
Among the adult Canthigaster at the RNP whose images are available from the 2022 expedition and iNaturalist 5 have only round spots on the body and head (Supplementary material: File S5A), 6 have radiating eye lines and a partial dorsal ocellus (Supplementary material: File S5B), and 17 have an intermediate color pattern, with scattered short lines and irregularities in spot form around the eyes and a varying degree of development of the dorsal ocellus (Supplementary material: File S5C, D). ODD and ETH (unpublished data) have found no genetic differences in terms of distribution of CO1 haplotypes within a haplotype network that included TEP fish with (n = 29) and without (n = 46) radiating eye lines (including 2 with and 6 without eye lines from the RNP) and individuals of C janthinoptera from the Indo-West Pacific. Sharing of CO1 lineages is present in various species complexes within Canthigaster (ODD, unpublished data). Given the blurred morphological character boundaries between species within this complex we include both species as part of the RNP fauna, where C. janthinoptera has not been recorded previously. We class both as residents because both are relatively common. We have not examined museum specimens from the RNP listed as “C. punctatissima” to determine how many might have color patterns consistent with C. janthinoptera.
Caranx caballus. This common, widespread TEP endemic carangid is known from southern California to Chile and all the offshore islands. It is a resident in the Galápagos (Victor, Grove et al., 2024), but probably a vagrant at Clipperton (Allen & Robertson, 1997). There is a long history of collections, in small numbers, and observations at all the RNP islands. However, in 2022 we photographed only 3 singleton individuals, ODD’s group collected 1 specimen it between 2015-2023 and there are only 3 separate images in iNaturalist, from 2021, including 2 images of a small group of fish in a multispecies group associated with a floating refrigerator ~ 30km offshore from San Benedicto (Green Jack flotsam). An SIO specimen collected in 1962 was taken in open water several miles from Socorro. This species is notable for repeatedly recruiting in large numbers of adults at the Hawaiian Islands, where it does not persist and is regarded as a non-resident (Mundy, 2005). Adults of this species commonly associate with offshore flotsam (Hunter & Mitchell, 1966, 1968) and have been repeatedly videoed at near-surface BRUVS deployed over deep water between Cocos and the Galápagos, far from land (Cambra et al., 2021; Vaudo et al., 2023). Caranx caballus, which ranges down to depths of 100 m, was not recorded in the ROV mesophotic study at the RNP islands by Hollarsmith et al. (2020) and Velasco- Lozano et al. (2020). Given how few of this species have been seen recently in the RNP it is possible that the low numbers there are maintained entirely by intermittent recruitment of adults from the mainland. Due to its persistently low abundance and strong potential for oceanic dispersal by adults we class this species as an uncertain resident.
Caranx caninus. There are no museum records of this species in the RNP. The only primary source records for this species are observations of single individuals at Roca Partida and San Benedicto in 2012 by coauthor AAB. Hence, we include it as a non-resident. C. caninus appears to be a specialist predator on engraulid fishes (Sánchez- García et al., 2017), none of which have been recorded from the RNP, a factor that may contribute to the lack of a population of this jack in the RNP.
Chaenopsis alepidota. The sole record of this species in the RNP is a photograph of part of the head of a single individual taken at Socorro in 2022. The identification is based on the color pattern of the head, which varies in a species-specific manner among members of that genus. This identification was confirmed by Dr. Phil Hastings, an expert in the taxonomy of TEP chaenopsids. Chaenopsis species are small, cryptic fishes found in sand and rubble habitats, which have not been extensively sampled in the RNP, particularly during this century. Hence, we record its population status as uncertain.
Chaetodon humeralis. The only museum records are from 1959 and 1961 collections at Socorro, housed at UBC and CMNFI (which was supplied with specimens from the UBC collection), the latter with a confirmed identification. There are iNaturalist images from Socorro taken in 2007 (a pair), 2020 (a pair) and 2024 (a single individual). ODD and students collected 4 adults from each of Socorro and Clarion in 2015, plus 2 more adults from Socorro in 2016, but none in 2019 or 2023. This species was not recorded during the 2022 expedition, or observed by AAB at Socorro, San Benedicto, and Roca Partida between 2012 and 2024. Hence, we consider it to be a non-resident, one that repeatedly arrives in the RNP in very low numbers.
Chriolepis sp. See endemics section B.
Cirrhitichthys oxycephalus. This Indo-Pacific species is widely distributed on reefs throughout the mainland (the Gulf of California to Peru) and all the offshore islands of the TEP. It is abundant at all 4 islands in the RNP, hence a resident. A recent genetic study spanning the longitudinal limits of its range (Red Sea to the Americas) found 2 major clades (Torres-García et al. 2024), one in the Indo-central Pacific and the other in the TEP. There are 2 subclades in the TEP, one at Clipperton and the other covering the rest of the TEP, with no indications that the RNP population is genetically isolated from the mainland population. That study also indicated that the name Cirrhitichthys corallicola Tee Van, 1940, type locality Colombia, could be resurrected for the TEP population if it is supported by a reexamination of morphological variation in this species (Torres-García et al., 2024).
Citharichthys sp. While the fish in the screen-grab image taken from a submersible video on sand at 60 m depth at Roca Partida in 2016 is identifiable as a species of Citharichthys (Supplementary material: File S2), it is not possible to determine which species it is due to the orientation of the fish in the image. The putative specimen of Citharichthys xanthostigma in USNM from Clarion was examined and is a misidentified Citharichthys platophrys. However, that record comes from the excluded 1889 Albatross collection compromised by problematic location data (Table 2). There are at least 6 species of this genus known from deep habitats around southern Baja, plus at least 2 insular endemic members of the genus and the Roca Partida fish could be any one of them. Hence, we include the Roca Partida fish as having an uncertain identity and population status.
Coryphopterus urospilus. See endemics section C.
Ctenochaetus marginatus. This surgeonfish, which is found in the central and eastern Pacific, is concentrated mainly from Costa Rica to Colombia on the mainland, with a few records in southern Mexico, and from all the oceanic islands of the TEP, including a multiple records at the Galápagos. Fourriére et al. (2016) recorded this species as observed in the late 1990s and 2007 (the latter by Chávez-Comparan et al. [2010], who did not record it as common), and, although they noted an SIO record from the RNP, the current online SIO catalog lists no specimens from there. There is only one other museum record, a UBC specimen collected in 1959 from Clarion. The few photographic records from 2022-2023 are almost entirely of solitary large adults, with the few iNaturalist images restricted to Socorro. ODD collected 2 individuals from Socorro and AAB has observed single adults and occasional pairs of adults at San Benedicto and Socorro. The general paucity of records, and recent observations restricted to adults of a readily recognizable surgeonfish indicates that C. marginatus has an uncertain residency status in the RNP. This species can live for at least 25 years and may recruit only intermittently from elsewhere to the RNP islands (Clements et al., 2012).
Dactyloscopus insulatus. See endemics section A.
Doryrhamphus melanopleura and Doryrhamphus paulus. See endemics section D.
Echidna nebulosa. This species is widespread and common in many parts of the TEP. There are single iNaturalist images from Socorro (2021) and Clarion (2023) and 3 different individuals were photographed at Clarion in 2022. ODD collected 1 specimen at Clarion in 2015 and 4 at Socorro in 2015, all of them in intertidal rock pools, which are now in CPUM. There are no published museum records from the RNP. Small, reef-living cryptic moray eels are secretive and may not be readily visible and there appear to be sufficient records to class this as a resident.
Echidna nocturna. This species is another small, cryptic moray that is widespread in the TEP. In the RNP there are museum collections from Socorro (CAS 1897 [2], 1925 [1], 1979 [1], SIO 1955 [1]) and Clarion (CAS 1897 [3], 1925 [3]). One individual was observed at San Benedicto in 2022, and ODD collected 9 specimens at Clarion in 2015 and 11 specimens at Socorro in 2015, 2016, and 2023 (all from intertidal rock pools) that are in CPUM. Due to repeated collections over more than 100 y and an abundance of recent collections we class this small, reef-living cryptic moray eel as a resident.
Enneanectes exsul. see endemics section A.
Epinephelus analogus. A single LACM lot of 3 specimens collected in 1971 was examined and found to be misidentified Epinephelus labriformis. A USNM specimen collected in 1938 (Schmidt & Schultz, 1940) was examined and confirmed to be E. analogus. This species also was included among those Chávez-Comparan et al. (2010) recorded as observed. Since there is only 1 confirmed record of this species in the RNP we class it as a waif without a resident population.
Epinephelus clippertonensis and Epinephelus labriformis. E. clippertonensis was described from Clipperton Atoll by Allen and Robertson (1999), who, at that time, thought it was a Clipperton endemic. That description involved comparisons of the morphology, including marked differences in coloration, of 312 E. clippertonensis from Clipperton and 80 E. labriformis from Mexico, Panama, and Peru. Among the latter were 7 USNM specimens from Clarion that were collected in 1889 by the Albatross expedition to the RNP islands and other parts of Mexico. Allen and Robertson (1999) identified those 7 as E. labriformis, based on morphological characteristics other than coloration, which would not have been clearly discernible in those more than century-old specimens. Specimens of some species from different Albatross voyages have been erroneously mixed with those from the 1889 RNP expedition (see comments in the section on unaccepted databases and species) and others have incomplete collection data. The Albatross E. labriformis specimens putatively from Clarion are part of the group with an incomplete collection date, which raises the possibility that they were collected elsewhere. Assessment of the morphology of the 2 species from our very recent collections in the RNP could help resolve the issue of the degree of resemblance of both forms to E. clippertonensis from Clipperton.
Since E. clippertonensis was described it has become evident that, in addition to Clipperton, fish with E. clippertonensis coloration also occur in the RNP (Del Moral-Flores et al., 2016; Fourriére et al., 2016; Hollarsmith et al., 2020; Robertson & Allen, 2006, 2015), southern Baja California, and Alijos Rocks, 290 km west of southern Baja California and 735 km north of the RNP (Craig et al., 2006; and see SIO specimen database). Review of images of live fish from the 2022 expedition and the few (12) available on iNaturalist in late 2024 shows the following, in terms of coloration: RNP fish include 126 E. clippertonensis, 6 E. labriformis and 3 with intermediate coloration (all of the latter 2 groups from 2022 images). Chávez-Comparan et al. (2010) mentioned only E. labriformis as present at Socorro, likely due to lack of knowledge about the differences between the 2 species. At mainland Mexico among images by some of the present coauthors, but mainly from iNaturalist, 168 have coloration of E. labriformis and 2 of E. clippertonensis, both at southern Baja. All images south of Mexico have E. labriformis coloration: Costa Rica (51), Panama (50), Cocos (7) and Galápagos (80). Thus, based on coloration, the great majority of the current RNP population comprises E.clippertonensis, with a few E. labriformis and possibly some hybrids, and E. clippertonensis is largely restricted to offshore islands.
Craig et al. (2006) examined genetic relationships, using CytB mtDNA, among 304 individuals of both species from the mainland between Baja California and the Gulf of California to Panama, plus Cocos, Galápagos, Clipperton and Alijos Rocks, but included none from the RNP. They found that all 24 individuals from Clipperton had E. clippertonensis coloration, but 20 had E. clippertonensis haplotypes and 4 had the most common mainland (E. labriformis) haplotype. While all but 1 of 12 individuals from Alijos Rocks had E. clippertonensis coloration, all but 1 also belonged to a haplogroup only present there, while the remaining individual had the most common E. labriformis haplotype. ODD (unpublished data) sequenced 14 individuals collected in the RNP (no information on coloration available) and they form 2 haplogroups, 11 in a haplogroup closely related to the unique Alijos Rocks haplogroup and 3 in a haplogroup most closely related to E. labriformis. These results support the conclusion that both species are in the RNP, although E. labriformis is much less common. However, the existence of a large, local haplogroup involving the great majority of individuals at the RNP (and other such haplogroups at Alijos Rocks and Clipperton) and a lack of a strong relationship between genetics and E. clippertonensis coloration among fish at Alijos Rocks, Clipperton, and possibly the RNP, indicates that the taxonomic status of E. clippertonensis is complex and currently unresolved. Craig et al. (2006) provided a detailed discussion of various possible explanations of the disconnections between coloration and genetics in E. clippertonensis. See also account of E. clippertonensis in endemics section D.
Forcipiger flavissimus and Forcipiger longirostris. Prior to 1898, when F. flavissimus was described from specimens collected in the RNP, fish from there were identified as F. longirostris, a name that carried over to some mid-20th century collections by UBC from the RNP. Three museums, CMFI, SIO and LACM, house 45 specimens from the RNP, including 19 individuals collected in 2022. The morphology of those was re- examined. Six from CMFI that were collected there in the 1950s and originally labelled F. longirostris are F. flavissimus. Twenty from SIO and 19 from LACM were correctly labelled F. flavissimus. ODD collected 15 individuals in the RNP, which have been examined morphologically and are F. flavissimus. Forty images of different individuals of Forcipiger in the RNP obtained during the 2022 expedition plus 11 from iNaturalist all appear to be F. flavissimus, based on snout length and, when visible, other features indicated in the images in Supplementary material: file S6. Similarly, all of the 25 identifiable (to species) images of Forcipiger taken elsewhere in the TEP that are available on iNaturalist plus 17 images of different individuals taken by CJE and AME at the Galápagos and Baja also appear to be F. flavissimus. Thus, there is no evidence that F. longirostris occurs in the RNP together with F. flavissimus, or anywhere else in the TEP, which is not included in the geographic range of F. longirostris in any modern publications that have such information (e.g., Allen, 1981; Mundy, 2005; Myers & Pratchett, 2010; Randall, 2005, 2007). Forcipiger flavissimus was 1 of the transpacific fishes examined by Lessios and Robertson (2006), who found that 5 individuals from Clipperton and 3 from the RNP had the most common sequence of ATPase mtDNA found among conspecifics from the western side of the Eastern Pacific Barrier, indicating a lack of long-term isolation of the TEP population.
Both species of Forcipiger are found throughout the ICP, indicating similar dispersal capabilities. Both are at all major island groups on the western border of the East Pacific Barrier, including the Line Islands, from where propagules of potential transpacific migrants can be dispatched towards the TEP on the equatorial countercurrent. In Hawaii, where both species are found, Randall (2007) noted that F. longirostris tends to occur in deeper water. The 2 also have dissimilar diets (Konow & Ferry-Graham, 2013; Randall, 2005), indicating that the presence of only 1 of them in the TEP likely is not due to competitive exclusion of the other. While there are no validated records of F. longirostris in the TEP close similarities in the appearance of those 2 species mean that attention should be given to careful review of images taken in the future to verify whether F. longirostris does arrive in the TEP on some occasion.
Gobiesox canidens. See endemics section A.
Grammonus diagrammus. This TEP endemic brotula is a highly cryptic black fish that lives deep in crevices and is rarely seen alive in its habitat. Such forms are rarely caught in abundance without the use of ichthyocides. The fact that it was caught in 2 different years, once with 5 different individuals, suggests that it has a resident population at Clarion.
Gymnothorax flavimarginatus. This is a large (to 120 cm), distinctively colored and not particularly secretive moray that is widely distributed in the Indo-Pacific. It is known from all the TEP oceanic islands, as a vagrant in the Galápagos (Victor, Grove et al., 2024), with a few records from the Mexican mainland and with most records concentrated in mainland Costa Rica and Panama. There are no museum records prior to its collection by ODD (specimens in CPUM) and it was not observed by Fourriére et al. in 1994-1999, 2007, 2010 or 2012-2013. Acceptance by the “two 2016 inventory” studies was based on secondary sources. A single specimen was photographed at Socorro in 2022 and a few reads of CO1 (100% name assignment) at Clarion during the 2023 eDNA study. This species is classed as an RNP non-resident.
Halichoeres adustus. See endemics section D.
Halichoeres insularis. See endemics section A.
Halichoeres nicholsi. See endemics section B.
Halichoeres sanchezi. See endemics section A.
Hemiramphus saltator. Although this species is listed in the 2002 checklist and both 2016 inventories, we were unable to locate any voucher items until a photograph of this species was taken in June 2025 by an iNaturalist contributor at Socorro (Supplementary material: File S4). That image showed a school of 4 individuals, which, the photographer stated, comprised no more than 10 individuals. Because of this paucity of records, we list this species as having an uncertain population status.
Heteropriacanthus carolinus. See Priacanthus alalaua.
Holacanthus clarionensis. See endemics section A.
Holacanthus passer. The only museum records of this species, which is common throughout the rest of the TEP, are a single specimen in CMNFI collected in 1968 at Socorro and 2 in CPUM, one collected at Clarion in 2014 and the other at Socorro in 2015. AAB recorded single individuals at San Benedicto in 2014 and another at Socorro in 2017. Fourriére et al (2016) reported observations in the late 1990s, 2010 and 2012-2013, but did not provide numbers of individuals or the islands at which they were seen. Hollarsmith et al. (2020) and Velasco-Lozano et al. (2020) did not record this species during their ROV mesophotic study. This species was not recorded in 2022. We include an image of a single individual photographed at Roca Partida in 2015, and an image of a likely hybrid of H. passer X H. clarionensis (Supplementary material: File S7) collected at Socorro during the 2022 expedition. Given the rarity of this conspicuous, readily recognizable, demersal species we regard it as a vagrant.
Hypanus dipterurus and Hypanus longus. While there are multiple validated records indicating that H. longus is a resident in the RNP there are many fewer of H. dipterurus. Due to morphological similarities of those 2 species some confusion is inevitable, since a large proportion of the much longer tail of H. longus, which is often used for identification, is frequently lost in live individuals. There are 2 putative RNP records of H. dipterurus in GBIF: one LACM specimen from Socorro, on re-examination, is H. longus and another from MNHN is from the Gulf of California, not the RNP. However, both species were accepted in the curated RNP inventory by Becerril-García et al. (2020), based on an observation at San Benedicto as well as a museum record of H. dipterurus from Socorro (likely the erroneous record from LACM). Since there is one confirmed image of H. dipterurus from Socorro (Supplementary material: File S4) its presence in the RNP is accepted as a vagrant due to the very small number of verified records.
Hyporthodus cifuentesi. Aburto-Oropeza et al. (2017) employed deep water BRUVs and recorded this species as abundant at multiple islands, including Socorro. Hollarsmith et al. (2020) videotaped it at Clarion and San Benedicto using a small ROV, and we include an image from each of those studies. Due to its abundance, we class it as a resident.
Hypsoblennius proteus. See endemics section A.
Iniistius pavo. This species is a medium-sized, Indo- Pacific wrasse found in shallow sand habitats around the fringes of reefs. It is widely distributed in the TEP, from central Pacific Baja to the Gulf of California and south to Colombia, as well as all the offshore islands except Clipperton (which has a minuscule amount of shallow sand habitat). It was photographed at 3 islands in 2022: an adult and a juvenile at each of Clarion and San Benedicto, and an adult at Socorro. AAB only observed it once, at Socorro, in 2023 and Fourriére et al. (2016) observed it in the 1990s, but not in 2007, 2010 or 2012-2013. There are no museum records of this species. It is classed as an uncertain resident due to this pattern of observations.
Kyphosus sectatrix and Kyphosus lutescens. See K. lutescens in unaccepted-species section.
Labrisomus socorroensis. See endemics section A.
Lutjanus argentiventris. This species, which is common on the TEP mainland, often seen in schools, is rare in the RNP. It has been repeatedly observed as single individuals at different locations in different islands in different years: at Clarion in 2013 and 2022 (Supplementary material: File S1), at San Benedicto in 2013, 2017, 2018, and 2024 (Supplementary material: File S3), and at Socorro in 2019. Its rarity and the lack of museum specimens indicates that it is a vagrant that repeatedly arrives from the mainland.
Lutjanus inermis. Chávez-Comparan et al. (2010) recorded this species on 50% of 37 transects at Socorro in 2007 and that it was the ninth most abundant species on those transects. Although there are no voucher specimens or photos of this species it seems reasonable to accept this sole group of observation records owing to the distinctive form and coloration of this species that make it readily distinguishable from other Lutjanus species. Local abundance in this species would not be surprising as it often occurs in large aggregations. As there are no other records of this species from the RNP, L. inermis is classed as a non-resident.
Lythrypnus cf. dalli. See endemics section B.
Lythrypnus insularis. See endemics section A.
Mobula tarapacana and Mobula thurstoni. The only record of M. tarapacana is an iNaturalist image in 2021 at Clarion (Supplementary material: File S1) and an observation at Roca Partida by Becerril-García et al. (2020). The only records of M. thurstoni are a video of a school of at least 25 individuals at Roca Partida in 2018, a photograph of a single individual at Clarion in 2022 and 6 reads of a CO1 ASV obtained by the 2023 eDNA sampling at San Benedicto and assigned to M. thurstoni with 100% confidence. G. Stevens (of Manta Trust, https://www. mantatrust.org/, pers. com. 2024) considers that both these species occur too intermittently to have a resident population in the RNP.
Mugil cephalus. The first visual record of this species in the RNP is from iNaturalist images taken by Alberto Alcalá at southern Socorro in May 2024 (Supplementary material: File S4), with only a small group (~ 6 individuals) observed then. He returned to the same site at Socorro in November 2024 and obtained a photograph of 27 individuals, which he said was part of a school of ~ 80 adults (Mugil). iNaturalist hosts other images of adults of this species taken at Socorro in 2021, 2023, and 2025. Despite the existence of a school of adults, we list residency as uncertain, since there are no other records of this species in the RNP and all those images seem to have been taken at the same general site at Socorro. The type locality for M. cephalus is in Europe. Global-scale genetic studies of M. cephalus, the most recent one by Thieme et al. (2025), have concluded that putatively pantropical M. cephalus actually comprises 16 allopatric species. Those include 2 species in the TEP, Mugil galapagensis (endemic to the Galápagos Islands) and another found from southern California to northern Peru. The only name of a species from that area that is listed in Eschmeyer’s Catalog of Fishes (Fricke et al., 2026) as a synonym of M.cephalus is Mugil mexicanus, type locality Acapulco, which likely is the name that should be applied to the Revillagigedo population as well as the mainland TEP population.
Mugil setosus and Mugil curema. Gilbert (1892) described M. setosus from Clarion, noting its similarity to M. curema and stating that it was abundant there (the type collection includes 14 specimens). M. setosus is 1 of 2 species names that have been proposed for the TEP member of the M. curema species complex. Durand and Borsa (2015), in a genetic study, hesitated to provide a name for TEP members of that complex, among others, until the TEP phylogeny was resolved and proposed the name “Mugil species O”. Their study revealed 2 close mtDNA lineages of that species on the mainland of the TEP. They did not have samples from the RNP in their analysis and did not mention M. setosus. New CO1 DNA- barcode data obtained by ODD (unpublished data) shows that all 6 sequenced RNP specimens belong to 1 of those 2 lineages, as do specimens from mainland Mexico and Costa Rica. The other lineage, 1.5% divergent from the former, occurs throughout the mainland of the TEP, from the Gulf of California to Ecuador. Britzke et al. (2019) compared the morphology of type specimens from Clarion and fish from the Gulf of California, Panama, Ecuador and Peru and found no morphological evidence for more than one species in the complex and concluded that M. sp. O is synonymous with M. setosus. Fourriére et al. (2016) listed only M. setosus for the RNP, while Del Moral-Flores et al. (2016) included both M. setosus and M. curema. Since the type location of M. setosus is the RNP and, to date, only a single lineage has been found there, we consider that to be M.setosus and that it is resident in the RNP.
Myrichthys pantostigmius. see endemics section D.
Myripristis clarionensis and Myripristis leiognathus. In color (uniform red) and form, these 2 species often look almost identical, so are easily confusable to the untrained eye or in less-than-ideal viewing circumstances. However, both species have large, obvious, well-defined scales on the body. The distinguishing feature visible to trained divers with a clear view of these fish and obvious in most well-focused images of these species is the number of scale rows between the lateral line and the spinous dorsal fin: 2.5 rows in M. leiognathus and 3.5 rows in M. clarionensis, with the half row running along the base of the dorsal fin above (and partly covered by) the row below of the 2 or 3 rows of fully visible scales above the lateral line (Supplementary material: File S8). All RNP museum specimens of these 2 species collected that have been examined to date (72 from SIO and 168 from LACM) have been called M. clarionensis, except for 2 juveniles and an adult collected in the 1960s that are housed in CMNFI and were re-examined and confirmed as M. leiognathus. ODD collected 21 specimens, all of which are M clarionensis. On the 2022 expedition we collected 8 M. clarionensis and 1 juvenile M. leiognathus and obtained identifiable images of 69 different individuals at the 4 islands, all of them M. clarionensis. iNaturalist has 7 identifiable images from the RNP labelled, correctly, as M. clarionensis, but none of M. leiognathus. From this we conclude, like Greenfield (1965), that, while M. clarionensis is a common RNP resident, M. leiognathus is an occasional vagrant there and does not have a resident population.
Myxodagnus opercularis. There are a number of museum collections of this species; at Clarion (LACM) in 1971, and at Socorro: 1 in 1925 (CAS), 3 individuals in 1934 (LACM), 4 in 1970 (SIO) and another 2 in 1971 (LACM). This fish is a very small, highly cryptic species that lives buried in the superficial layer of sand and hence unlikely to appear in collections that do not use ichthyocides, which have not been used since the mid-20th century to collect fishes in the RNP. Since multiple individuals were repeatedly collected at Socorro in each of 4 different years over a 46-year period and it was collected at the end of that period at 2 islands we class it as a resident.
Ophioblennius clippertonensis and Ophioblennius steindachneri. There are 2 currently recognized species of this genus in the TEP, O. steindachneri and O. clippertonensis. The latter was first described as a subspecies of O. steindachneri by Springer (1962) based on morphological differences and he included the RNP population, from which he examined 25 specimens collected in 1953 at Socorro (LACM 48933.007), as part of O. steindachneri steindachneri. Subsequently those 2 were raised to species status by Hastings and Springer (2009). To date, O. steindachneri has been thought to be the only species present on the TEP mainland and oceanic islands except Clipperton, which is occupied by the endemic O. clippertonensis (Fricke et al., 2026). Del Moral-Flores et al. (2016) and Fourriére et al. (2016) include only O. steindachneri as part of the RNP fauna. However, DNA barcodes by ODD (unpublished data) of 34 individuals from the Revillagigedos, 17 from Clipperton and 26 from the mainland and Galápagos show 3 haplogroups, 1 for mainland/Galápagos fish (O. steindachneri), separated by 1.3% of uncorrected pairwise genetic distances (p-distances) from 1 haplogroup for Clipperton and 1 haplogroup from the RNP that is closely related (p-distance = 0.6%) to the Clipperton haplogroup. In addition, a few O. steindachneri are present at the Revillagigedos (5 of 34 total) and Clipperton (3 of 17 total) and a single “clippertonensis” haplotype is present on mainland Mexico. Given the present information the simplest explanation is that both species, with a predominance of O. clippertonensis, currently are in the RNP, although further study, particularly involving comparisons of the morphology of fishes collected in 1953 with those collected 70 years later during the present study, may well change that conclusion. Images of Ophioblennius from the RNP show a similar range of variation in coloration to that seen in O. steindachner on the mainland and Galápagos.
Orthopristis cantharina. Once thought to be endemic to the Galápagos, this species is known from Baja California and the Gulf of California as well (GBIF-O. cantharina; Robertson & Allen, 2024). There are no reliable museum records from the RNP (see comments about the problematic databases, below). However, video by AAB at San Benedicto shows what most likely is that species, which has a distinctive shape. The screen-grab image (Supplementary material: File S3) shows 2 large individuals (the species reaches at least 45 cm TL) near a rock on sand adjacent to rocky reef. The more strongly colored individual in the background was swimming rapidly to-and-from around the other individual. Based on this single observation we class O. cantharina as a non-resident.
Pareques sp. See endemics section B.
Platax teira. The native range of this species is restricted to the Indo-West Pacific, with its nearest population about 7,000 km from Mexico. In 2023 a single large adult was photographed at Cabo Pulmo, southeastern Baja, the first record of this species in the TEP. Subsequently, in 2024 and 2025, juveniles and multiple adults have been photographed at various sites scattered along 1,000 km of the southern Mexican mainland coast (Medina-Rosas & Moreno-López, 2025; Petatán-Ramírez et al., 2025). An adult was photographed on the west coast of Socorro in May 2025 and another on the east coast in June 2025 (Supplementary material: File S4). Subsequently an adult was photographed at Roca Partida in December 2025 (Supplementary material: File S2). Because these single, solitary adults provide the only records to date of this species in the RNP we include it as a non-resident. The sudden appearance of adults at very small, isolated locations such as Roca Partida indicates that they, like juveniles, can disperse pelagically, most likely associated with flotsam.
Priacanthus alalaua and Heteropriacanthus carolinus. Priacanthus alalaua is known from the Hawaiian Islands, Guam and Japan as well as the RNP, plus Alijos Rocks, Guadalupe and southwestern Baja California (Hashimoto & Motomura, 2025; Starnes, 1988). Fish in the RNP, where it is a resident, reach a larger size and occur in shallower water than this species at Hawaii (Fitch & Crooke, 1984; Starnes, 1988). Heteropriacanthus carolinus is an Indo-Pacific priacanthid found throughout most of the TEP, including all the oceanic islands, and is a common resident in shallow water in the RNP. The similarity in general body shape and reddish coloration of these 2 species likely will make them difficult to distinguish in underwater images of live individuals that are not high quality closeups. Morphological differences in the preoperculum that distinguish between these 2 species are indicated in File S9.
Pristigenys serrula. The occurrence of this species at the RNP derives from 3 sources. Starnes (1988) noted it as “recorded” there but provided no more information. Del Moral-Flores et al. (2016) and Fourriére et al. (2016) cited secondary sources (Castro-Aguirre & Balart, 2002; Robertson & Allen, 2006, 2015), that likely were based on Starnes (1988). Fourriére et al. (2016) also referred to an occurrence in FishBase, an MNHN record from a problematic database that has been entirely excluded (see comments below about unaccepted databases). Hollarsmith et al. (2020) recorded 2 observations of single individuals at depths of 76-84 m at Clarion, but none at San Benedicto or Socorro. We have included a screen- grab image from that study at Clarion (Supplementary material: File S1). Although these are the only records of this species it ranges down to 250 m depth and is usually found at depths below scuba activity and likely is an RNP resident.
Prognathodes falcifer. This species has a highly distinctive form and color pattern that differs from all other chaetodontids in the northern TEP and allows ready recognition. Fourriére et al. (2016) recorded observations between 1994-1998 and 2010. Hollarsmith et al. (2020) recorded it 12 times at Clarion, San Benedicto and Socorro in 2018 and we include 2 screen-grab images from videos used in that study (Supplementary material: Files S1, S4). Video taken by Ayala-Bocos et al. (2015) at Roca Partida and San Benedicto (Supplementary material: Files S2, S3) shows it to be common at depth. Hence, we regard it as a resident.
Pronotogrammus multifasciatus. This deep-reef species (depth range 40-400 m) is known from California to northern Peru and all the oceanic islands except Clipperton. The only museum specimen was a juvenile collected in 1897, which has been lost from CAS. The UNESCO eDNA project that sampled in 2023 recorded several thousand reads of 2 ASVs of 16S (100% confidence) at each of Clarion, San Benedicto and Socorro. We accept the reliability of the species assingnment for that gene because P. multifasciatus is the only member of its genus in the TEP. Due to the abundance of such reads of a species unlikely to be accessible to shallow collecting at the 3 major islands we class it as a resident.
Rypticus courtenayi. See endemics section A.
Scomber australasicus and Scomber japonicus. Scomber australasicus, which is genetically distinct from S. japonicus (Catanese et al., 2010) evidently has a resident population in the RNP with numerous individuals sampled there over more than 2 decades. There are also a few records in the southwestern Gulf of California. A global phylogeny of Scomber based on mtDNA by Scoles et al. (1998) showed no differentiation of the RNP population from those elsewhere in the Pacific, indicating that the RNP population is unlikely to be endemic. In contrast, the only RNP records available for S. japonicus are CAS specimens collected in 1925, 3 from Clarion and 1 from Socorro. The fact that so few S. japonicus were collected almost 100 years ago, and none since then, indicates it is not resident in the RNP.
Scorpaena afuerae. The 2 records of this deep-reef species (depth range 35-100 m) in the RNP are from videos obtained from a submersible at Roca Partida by Ayala- Bocos et al. (2015) and by an ROV at Socorro by CS-O in 2016. We have identified these as S. afuerae based on those fish having a red body with their pelvic fins and the inner surface of their pectoral fins being white with large red spots (Supplementary material: Files S2, S4). This color pattern is only seen in S. afuerae among all the known species of Scorpaena in the TEP. We tentatively class this species as a resident, based on these 2 records.
Scorpaenodes xyris. This small reef-dwelling scorpaenid, which was described from Baja specimens by Jordan and Gilbert (1892), is common at all 4 RNP islands. A genetic study by Bernal-Hernández et al. (2024) found that while all 44 individuals from the RNP belong to a local mitochondrial lineage, 3 fish from that lineage occur in the Cortés province, 2 in the Panamic province (central Mexico to Peru), and 2 at Clipperton Island. As a result, the RNP population has only a small degree of isolation (0.3% separation) from the other populations.
Selene peruviana. There are a few old records of this benthopelagic, conspicuous and readily recognizable species: 1 from Socorro in 1959, now at CMNFI, which confirmed that identification, another from Socorro at collected in 1972 SIO (collection location uncertain) and 4 at USNM (Albatross 1889 specimens, confirmed identification, excluded from use due to incomplete collection data). We class this species as an RNP non-resident.
Seriola dorsalis. Seriola dorsalis, a large, conspicuous, readily recognizable jack, is a north Pacific species recently split from what was once regarded as a globally distributed species, Seriola lalandi (Martinez-Takeshita et al., 2015). Before the 2023 eDNA study the only primary-source record was a 2012-2013 observation listed by Fourriére et al. (2016). That eDNA study produced a very small number of reads of CO1 and several thousand reads of 16S from the 3 main RNP islands, with 100% assignment to S. lalandi. Based on those genetic data we record it as present, but without a resident population.
Serranus aequidens. Serranus aequidens is a deep- living species (depth range 75-500 m) that would be unlikely to appear in shallow water collections. In addition to 2 USNM specimens collected at Socorro in 1934, whose identity was confirmed by our re-examination, a small number of reads of CO1 with 100% assignment to S.aequidens were obtained at Clarion by the 2023 eDNA research. At present we regard its population status in the RNP as uncertain.
Serranus socorroensis. See endemics section A.
Stegastes flavilatus, Stegastes rectifraenum and Stegastes acapulcoensis. The 5 species of Stegastes for which there are museum collection records in the RNP can be distinguished by various combinations of the number of soft rays of the dorsal, anal and pectoral fins plus the number of lower gill-rakers on the first gill arch, as well as species-specific coloration of live and recently collected individuals (Allen & Woods, 1980). Jordan and McGregor (1899), who made the earliest collections of Stegastes in the RNP, from Clarion, San Benedicto and Socorro, identified them as S. flavilatus and S. rectifraenum. Subsequently, Heller and Snodgrass (1903) included those authors specimens of both species as Stegastes redemptus when they described that species. Ricker (1959) noted that those preserved specimens lacked color that could have been useful for identifying them.
GBIF contains information about UBC specimens named S. flavilatus and S. rectifraenum that were made at Clarion and Socorro between 1957-1963 by P.A. Larkin and colleagues. Those include multiple lots involving numerous individuals of both entities, some of which were transferred to CMNFI. Our examination of 13 individuals in the 1957-1963 CMNFI collection labelled S. flavilatus shows that all were misidentified S. redemptus. In addition, our examination of 20 individuals in 6 lots from Clarion and Socorro between 1959-1963 that were labelled S. rectifraenum shows that 18 were correctly identified and 2 were misidentified Stegastes leucorus. Those UBC collections produced only 1 individual named S. redemptus from Clarion in 1963, and 2 lots of 3 individuals from Socorro in 1962-1963, i.e., at the end of that multi- year collecting effort. However, 2 UBC lots comprising 35 individuals of S. redemptus, but including no S. flavilatus or S. rectifraenum, were obtained from Socorro in 1954 by a different collector (M.A. Newman). We confirmed the identification of 22 of those S. redemptus from 4 collection lots. Collections made between 1925-1955, that are deposited in other museums, also include numerous S. redemptus from both those islands, but no S. flavilatus or S.rectifraenum: 4 lots at LACM and FMNH collected at Clarion, together with 48 lots of 82 individuals, including 1 lot of 32 fish, at CAS. Collections between 1925 and 1955 of S. redemptus, but no S. flavilatus or S. rectifraenum, at Socorro include 17 individuals at CAS, 29 individuals in 3 lots at LACM in 1971-1973 and 3 individuals at SIO during 1956-1973. Thus, collecting efforts by different museums at different times in the mid-20th century at the 2 largest islands in the RNP produced very different numbers of S.rectifraenum as well as an abundance of S. redemptus. Castañeda-Beltrán (1988) collected 9 individuals he identified as S. rectifraenum from Clarion, and visually recorded many individuals under that name as well but made no mention of observing or collecting either S. flavilatus or S. redemptus, which suggests those fish were misidentified. There is a 2007 observation record of S. rectifraenum in Fourriére et al. (2016), but no information about what life stage it involved. Those authors also list LACM as a source of specimens of S. rectifraenum. However, there are no specimens of that species from the RNP in the LACM catalog.
Since the mid-20th century the only vouchered records of S. flavilatus have been an iNaturalist image of a juvenile S. flavilatus at Socorro in 2007, our observations of multiple small juveniles (but no adults or large juveniles) of S. flavilatus at San Benedicto in 2022 and our collection of a juvenile S. flavilatus at Clarion in 2022. We also observed multiple small juveniles of what could have been either S. acapulcoensis or S. rectifraenum (small juveniles of these 2 species have identical coloration) at San Benedicto. DNA barcoding produced 100% matches of the sequences of 2 of those that we collected to S. acapulcoensis, although we recognize that S. rectifraenum and S. acapulcoensis are closely related and broadly share an mtDNA lineage. In contrast, in 2022 we observed and photographed many S.redemptus ranging in size from juveniles to large adults at Clarion, San Benedicto, and Socorro. Given the known misidentification of S. flavilatus from mid-20th century collecting events and that adults of only S. redemptus and S. leucorus have been collected and observed over more than 50 y, we conclude that the only resident Stegastes species in the RNP are those 2 species, which are both common there (Supplementary material: Files S1-S4). However, juvenile S. flavilatus and S. acapulcoensis occasionally recruit there from the mainland but do not establish resident populations. We class S. acapulcoensis, S.flavilatus, and S. rectifraenum as non-residents.
Stegastes leucorus. See endemics section D.
Stegastes redemptus. See endemics section A.
Stethojulis bandanensis. This transpacific wrasse is the only member of its genus in the TEP. ODD collected 2 individuals of this species at Socorro in 2015 and six at Clarion in 2016, which are housed at CPUM. These represent the only museum records. Single individuals were photographed at San Benedicto and Socorro in 2022, and there are iNaturalist images of 4 single individuals taken at Socorro in 2007, 2023, 2024, and 2025. The 2023 eDNA expedition produced 51 reads of 2 CO1 ASVs (99- 100% name assignment confidence) at Socorro and San Benedicto. Because this species has no records prior to 2007 and so few individuals have been observed, we do not regard it as having a resident population.
Thalassoma virens. See endemics section D.
Tomicodon absitus and Tomicodon sp. See endemics sections A and B, respectively.
Tylosurus fodiator. This species, which is endemic to the TEP, extends from the Gulf of California to Ecuador and all the offshore islands, except Clipperton. While there are RNP museum records from the early to mid-20th century, the numbers of individuals involved are low. A single individual was seen and photographed during the 2022 expedition and AAB recorded none at any island between 2012-2024. There are iNaturalist images from 2020, 2021, 2024, and 2025 at Socorro, and one from 2024 at Clarion. However, most such images are of fish at a distance, making certain identification difficult. It is large, conspicuous and easy to catch by hook and line. The 2023 eDNA sampling produced 66 reads of CO1 (100% species assignment to Tylosurus crocodilus (of which T. fodiator was, until recently, classed as a subspecies) at Clarion and San Benedicto. Due to the combination of the paucity of collecting and photographic records and the eDNA data we class it as an uncertain resident in the RNP.
Tylosurus melanotus. This Indo-Pacific species, another large member of the genus, is known from the Gulf of California to Colombia, plus all the offshore islands except the Galápagos. Collette and Banford (2001) stated that T. melanotus replaces Tylosurus pacificus at the Revillagigedos and other offshore islands in the TEP, although only T. pacificus is known from the Galápagos as a vagrant (Victor, Grove et al., 2024). There are 2 RNP collection records of single individuals collected at Socorro. The 1952 specimen initially cataloged in USNM as T. pacificus was examined by Collette and Banford (2001) and reidentified as T. melanotus. LACM also has a specimen identified as T. pacificus, collected at Socorro in 1955, which our re-examination showed to be a misidentified T. melanotus. Del Moral-Flores et al. (2016) and Fourriére et al. (2016) recorded both T. pacificus and T. melanotus at the RNP, citing both LACM and USNM as sources of the T. pacificus record but not providing a primary source for T. melanotus. The 2023 eDNA sampling produced 64 reads of CO1 with 100% species assignment to Tylosurus acus, of which T. melanotus was previously classed as a subspecies, at all 4 islands. Due to the paucity of (old) museum records combined with the few recent eDNA data, we class it as an uncertain resident in the RNP.
Xyrichtys sp. See endemics section B.
Unaccepted databases and species
Two rejected databases: i) a database at the USNM from the Albatross 1889 Expedition. In February-April 1889 the USS Albatross collected fishes in Pacific Mexico, including Guadalupe Island, Alijos Rocks, the Revillagigedo Islands, the Gulf of California and the Pacific coast of Baja California (Tanner, 1889). Fishes were sampled at 3 Revillagigedo islands between March 4-11, Clarion for 1.25 days, then Socorro for 2.5 days and finally San Benedicto for less than 1 day. Specimens from that expedition are in the collections of the USNM and MCZ. The MCZ collection is relatively small, with no obvious errors. However, the USNM collection was evidently assembled in pieces at different times, as indicated by large variation in catalog numbers, and various records have vague collection dates. The 131-series of catalog numbers in particular has a relatively large number of location errors and vague collection dates (no date, or 1889 without a day and month date, or March 1889 rather than the known collection dates at the RNP islands). That 131-series also includes 3 ariid catfishes supposedly obtained at Clarion: Ariopsis sp., Notarius insculptus, and Notarius troschelii. Members of this family are egg- brooders, do not have a pelagic juvenile phase that would enable oceanic dispersal and lack confirmed records from any of the TEP offshore islands. Further, N. insculptus is restricted to Panama and Colombia and the specimens of these 3 species apparently came from a single collection in Panama or Colombia, which were visited by the Albatross on a different cruise to that at the RNP. Because ariids are not reef-associated fishes in the TEP they are not included among the unaccepted group of such species considered here. This case provides an example of the degree of unreliability of the 131-series dataset, which is why we excluded from the accepted species list all species for which 131-series records represent the sole RNP record: Chaetodipterus zonatus, Chloroscombrus orqueta, C. platophrys, Eucinostomus argenteus, and Symphurus leeorum. The 131-series records also include Diplectrum euryplectrum, which, since the LACM 1939 record of that species is also erroneous, is also excluded from the accepted species list. In addition, we also discounted 131-series records for Caranx sexfasciatus, Caulolatilus princeps, O. cantharina, and S. peruviana, all of which have more recent records that we regard as validated and, hence, retain them as members of the RNP fauna. Based on known or likely erroneous location records from the Albatross USNM collection for species with various non- 131-series catalog numbers we also excluded Orthopristis chalceus (known from mainland Mexico), Mugil thoburni (a Galápagos endemic; Durand & Borsa, 2015; Victor, Grove et al., 2024), Centropyge bicolor/= Holacanthus bicolor/= Holacanthus mesoleucus (an Indo-West Pacific species recorded in the USNM ledger as from St Lucia), Sebastes carnatus (a temperate NE Pacific species), S. peruviana (identification confirmed, but incomplete collection data) and Guentheridia formosa (from Panama in the ledger; a species restricted to the mainland between El Salvador and Ecuador). Finally, 2 USNM Albatross- collection specimens labelled Carcharhinus plumbeus that we examined are both newborn size and lack complete data on collection dates. Because Carcharhinus species are notoriously difficult to identify at that stage of development and there are no other validated records of this species in the TEP the RNP record of this species is not accepted. Location errors are not uncommon in 19th century and early 20th century collections from expeditions that visited multiple, widely dispersed sites, which were then divided between collections in different institutions then, much later, reassembled in one museum. These including errors related to the Galápagos fish fauna arising from the 1888 Albatross expedition to the Galápagos, Colombia and Panama (Victor, Grove et al., 2024). ii) A Muséum National d’Histoire Naturelle, Paris (MNHN) database. There is 1 database in GBIF that we excluded entirely, which is from the Muséum National d’Histoire Naturelle, Paris (MNHN). This, which is geolocated in GBIF at 20o N latitude, -110o W longitude, at the northeastern corner of the RNP, has been cited as support for some fish occurrences in the RNP by Del Moral-Flores et al. (2016) and Fourriére et al. (2016). It is a compendium of records of 71 species with collection dates between 1867 and 1999 that, in GBIF, includes the statement that it is from the “Baie de Californie”, i.e., the Gulf of California. For some species listed in that database collecting locations in that Gulf are actually included, e.g., Espíritu Santo Island for Syacium ovale. It also includes records of west Atlantic and Indo-west Pacific species. Incorrectly georeferencing alone makes it irrelevant to RNP occurrences. That database includes records we reject for 15 accepted species and 6 unaccepted species (Tables 2, Supplementary material: Table S1). Assessments leading to non-acceptance of species in the RNP fauna Our assessments of the 364 reef-associated fishes that potentially are members of the RNP fauna indicate that there are erroneous records of 121 species in 87 genera, 54 families, eliminating 33% of the assessed species, 19% of the assessed genera and 16% of the assessed families (Table 2). Of those 121 species, 46 were misidentified or were unidentifiable and 4 were synonymized with other species also included by both 2016 inventories. Further, 37 lacked vouchers, acceptance of their records in the “two 2016 inventories” having been based on secondary sources (Bautista-Romero et al., 1994; Castro-Aguirre & Balart, 2002; Robertson & Allen, 2006, 2015) and lacked verifiable primary-source data (specimens, images or reliable observations). For 45 species review of location data showed that their records were not from within the RNP, while in 22 species all specimens were lost, permanently precluding assessment and 2 species were represented solely by pelagic larvae collected well offshore, and hence were not accepted. Finally, 40 of the 121 species had multiple issues relating to their rejection, e.g., misidentification plus unreliable location data. Of those 121 species 32 were not mentioned by either of the “two 2016 inventories”, 50 had been accepted by both and another 11 accepted on one inventory but not mentioned by the other. In addition, 8 species were accepted on one inventory but rejected on the other and only 3 species had been rejected in both inventories. Re-examination of museum specimens during the assessment process established that in 45 cases, involving 28 species from 8 museums, those specimens were misidentified accepted species.
| Family | Recorded identification | Name authority | Correct identification | Fourr. | DMF | Problems | Data and sources relating to problems |
|---|---|---|---|---|---|---|---|
| Acanthuridae | Acanthurus achilles# | Shaw, 1803 | – | Yes | ID, voucher | Chan74: grey literature, many identification issues | |
| Acanthuridae | Prionurus punctatus | Gill, 1862 | P.laticlavius | Yes | Yes | synonym | Ld19: synonym; both species recorded by Fourr & DMF |
| Aetobatidae | Aetobatus narinari | Gill, 1865 | Yes | Yes | voucher | BG20: no evidence; TEP form now named A. laticeps | |
| Anthiadidae | Acanthistius pictus | (Tschudi, 1846) | – | – | lost, location | UBC: specimen lost; Peruvian species | |
| Apogonidae | Apogon dovii | Günther, 1861 | Yes | Yes | voucher, location | No primary source; southern TEP species | |
| Apogonidae | Apogon guadalupensis | Osborn & Nichols, 1917 | A.atricaudus | Yes | Yes | synonym | Lea22: synonym; both species recorded by Fourr & DMF |
| Atherinopsidae | Atherinops affinis | (Ayres, 1860) | Doubt | – | voucher | No primary source | |
| Balistidae | Melichthys vidua | (Richardson, 1845) | Yes | Yes | ID | DMF: cite Rick, which does not mention M. vidua | |
| Balistidae | Sufflamen fraenatum# | (Latreille, 1804) | S.verres# | – | Reject | ID | J&M; USNM* |
| Balistidae | Xanthichthys auromarginatus# | (Bennett, 1832) | Sufflamen verres# | – | – | ID | Giddens: DRR reviewed video to identify* |
| Balistidae | Xanthichthys lineopunctatus# | (Hollard, 1854) | X. mento# | – | Yes | ID | SIO*; CUMV*; MCZ*; CAS (lost); UBC |
| Balistidae | Xanthichthys ringens | (Linnaeus, 1758) | X.mento# | – | – | ID, location | CMNFI*; Atlantic species |
| Belonidae | Tylosurus pacificus# | (Steindachner, 1876) | T.melanotus# | Yes | Yes | ID | USNM*; LACM*; C&B2001: ID* |
| Carangidae | Chloroscombrus orqueta# | Jordan & Gilbert, 1883 | Yes | Yes | location | Albatross 1889: USNM*, unreliable location | |
| Carangidae | Decapterus punctatus | (Cuvier, 1829) | D. muroadsi | – | Reject | ID, location | CMNFI*; Atlantic species; DMF: inferred is D. muroadsi |
| Carangidae | Euprepocaranx dorsalis | (Gill, 1863) | Alectes ciliaris# | Yes | Yes | ID | LACM*; E. dorsalis is name change from Carangoides otrynter |
| Carangidae | Gnathanodon speciosus | (Forsskål, 1775) | – | Yes | voucher | No primary source | |
| Carangidae | Hemicaranx zelotes | Gilbert, 1898 | Yes | – | voucher | No primary source | |
| Carcharhinidae | Carcharhinus acronotus | (Poey, 1860) | – | – | lost, location | UBC: lost; Atlantic species | |
| Carcharhinidae | Carcharhinus brachyurus | (Günther, 1870) | C.falciformis | Yes | Yes | ID, lost | LACM*; CSLB*; UBC: specimen lost; BG20: not mentioned |
| Carcharhinidae | Carcharhinus leucas# | (Muller & Henle, 1839) | Yes | Yes | ID | CB88: description vague; BG20: not accepted | |
| Carcharhinidae | Carcharhinus melanopterus# | (Quoy & Gaimard, 1824) | – | – | lost | UBC: specimen lost; present at Isla del Coco | |
| Carcharhinidae | Carcharhinus perezii | (Poey, 1876) | – | – | lost, location | UBC: specimen lost; Atlantic species | |
| Carcharhinidae | Carcharhinus plumbeus# | (Nardo, 1827) | Doubt | – | ID | B-G: unconfirmed; Albatross 1889: USNM – identification unreliable | |
| Chaenopsidae | Acanthemblemaria crockeri# | Beebe & Tee- Van, 1938 | A.mangognatha*# | – | – | update | SIO*; UBC*; records precede description of A. mangognatha |
| Chaenopsidae | Acanthemblemaria hancock#i | Myers & Reid, 1936 | A.mangognatha*# | Doubt | – | update, location | Rick; H&R*; southern TEP species; records precede description of A. mangognatha |
| Chaenopsidae | Acanthemblemaria macrospilus$ | Brock, 1940 | A.mangognatha*# | Doubt | – | update | CAS*; SIO*; records precede description of A. mangognatha |
| Chaetodontidae | Chaetodon meyeri | Bloch & Schneider, 1801 | Yes | Yes | voucher | No primary source | |
| Chaetodontidae | Forcipiger longirostris# | (Brousonnet, 1782) | F.flavissimus# | Yes | Reject | ID | CMNFI*; SIO*; CAS |
| Cirrhitidae | Oxycirrhites typus | Bleeker, 1857 | Yes | Yes | voucher | No primary source | |
| Clinidae | Heterclinus tristis | (Klunzinger, 1872) | Axoclinus multicinctus# | – | – | ID, location | UBC*: genus endemic to Southwest Pacific correct Rosenblatt; |
| Congridae | Ariosoma gilberti | (Ogilby, 1898) | A. hemiaspidus | Yes | Yes | larvae | SIO*: larvae |
| Congridae | Heteroconger digueti | (Pellegrin, 1923) | Yes | Yes | lost | CAS: lost; no other primary sources | |
| Cyclopsettidae | Citharichthys xanthostigma | Gilbert, 1891 | C. platophrys# | Doubt | Yes | ID, location | Albatross 1889: USNM*, location unclear |
| Cyclopsettidae | Citharichthys gilberti | Jenkins & Evermann, 1889 | Citharichthys sp.# | Yes | Yes | ID, larvae | SIO*: larvae collected well offshore, fin-ray counts not of C. gilberti; LACM*: small, unidentifiable to species |
| Cyclopsettidae | Syacium latifrons | (Jordan & Gilbert, 1882) | Doubt | – | voucher | No primary source | |
| Cynoglossidae | Symphurus leeorum | Jordan & Bollman, 1890 | Yes | – | location | USNM: location doubtful | |
| Diodontidae | Diodon eydouxii | Brissout de Barneville, 1846 | Yes | Yes | voucher | No primary source | |
| Dussumieridae | Etrumeus acuminatus | Gilbert, 1890 | Yes | – | voucher | B-R94: grey literature, unconfirmed | |
| Echeneidae | Remora albescens | (Temminck & Schlegel, 1850) | Yes | Yes | lost, voucher | CAS: lost; no other primary sources | |
| Embiotocidae | Embiotoca jacksoni | Agassiz, 1853 | Doubt | – | location | CAS*: San Benito (Baja), not San Benedicto | |
| Epinephelidae | Alphestes multiguttatus | (Günther, 1867) | Yes | Yes | voucher | CAB; no primary source | |
| Epinephelidae | Epinephelus quinquefasciatus | (Bocourt, 1868) | Yes | Yes | lost, ID, voucher | SIO*: lost & too small to reliably identify to species; no voucher | |
| Epinephelidae | Hyporthodus niphobles | Gilbert & Starks in Gilbert, 1897 | Yes | Yes | location | SIO: outside RNP | |
| Epinephelidae | Mycteroperca jordani | (Jenkins & Evermann, 1889) | Yes | Yes | voucher | No primary source | |
| Epinephelidae | Mycteroperca prionura | Rosenblatt & Zahuranec, 1967 | Epinephelus labriformis# | Yes | – | ID | LACM* |
| Epinephelidae | Mycteroperca sp. | E.labriformis# | – | – | ID | LACM* | |
| Ephippidae | Chaetodipterus zonatus# | (Girard, 1858) | Yes | Yes | location | Albatross 1889: USNM*, ID correct but location unclear | |
| Fistulariidae | Fistularia petimba | Lacepede, 1803 | F.commersonii | – | – | ID, location | LACM*; UBC; DMF suggest F.corneta; pantropical except for TEP |
| Gerreidae | Diapterus brevirostris | Sauvage, 1879 | Yes | Yes | voucher | no primary source | |
| Gerreidae | Diapterus sp. | – | – | ID, location | USNM*: unidentifiable, location suspect | ||
| Gerreidae | Eucinostomus argenteus# | Baird & Girard, 1855 | – | – | location | Albatross 1889: USNM, location Panama; west Atlantic species | |
| Gerreidae | Eucinostomus gracilis | (Gill, 1862) | – | – | location | UBC: location = Tres Marías Is. | |
| Ginglyostomidae | Ginglymostoma unami | Del Moral- Flores et al., 2015 | – | – | voucher | BG20: no evidence; MNHN: Gulf of California; no primary source | |
| Girellidae | Girella nigricans | (Ayres, 1860) | Yes | Yes | lost, voucher | CAS: lost; no other primary sources | |
| Gobiesocidae | Arcos sp. | Tomicodon sp.# | – | – | ID | UBC*: small, unidentifiable to species | |
| Gobiesocidae | Gobiesox aethus# | (Briggs, 1951) | G. canidens# | Yes | Yes | ID | ODDDNA: only G. canidens present |
| Gobiesocidae | Gobiesox adustus# | Jordan & Gilbert, 1882 | G.canidens# | Yes | Yes | ID | ODDDNA: only G. canidens present |
| Gobiesocidae | Tomicodon eos | (Jordan & Gilbert, 1882) | Yes | Reject | voucher | SIO: not in catalog; no primary source | |
| Gobiesocidae | Tomicodon zebra | (Jordan & Gilbert, 1882) | Yes | Yes | voucher | SIO: not in catalog; no primary source | |
| Gobiidae | Barbulifer pantherinus | (Pellegrin, 1901) | – | Reject | location | MNHN: Gulf of California; no primary source | |
| Gobiidae | Lythrypnus pulchellus# | Ginsburg, 1938 | L. insularis# | Yes | Yes | update | Records precede L. insularis description |
| Gobiidae | Lythrypnus rhizophora# | (Heller & Snodgrass, 1903) | L.insularis# | Doubt | reject | update, location | Records precede L. insularis description; L. rhizophora is southern TEP species |
| Gobiidae | Lythrypnus zebra# | (Gilbert, 1890) | L. insularis# | Yes | Yes | update | Records precede L. insularis description |
| Grammistidae | Rypticus bicolor | Valenciennes, 1846 | R.courtenayi# | Doubt | Reject | ID, lost | CMNFI*; SIO*; CAS: lost |
| Grammistidae | Rypticus saponaceus | (Bloch & Schneider, 1801) | R.courtenayi#? | – | Reject | location | DMF: misidentified R. courtenayi; Atlantic species; MNHN: Gulf of California |
| Haemulidae | Anisotremus surinamensis | (Bloch, 1791) | A. perezponcedeleoni# | – | – | ID, location | UBC*; West Atlantic species |
| Haemulidae | Orthopristis chalcea | (Günther, 1864) | Yes | Reject | location | Albatross 1889: USNM*, unreliable location | |
| Kyphosidae | Kyphosus lutescens# | (Jordan & Gilbert, 1882) | K.sectatrix# | – | Yes | synonym | K&C13,16: synonym; ODDDNA supports synonym |
| Labridae | Bodianus pulcher | (Ayres, 1854) | – | – | lost, location | CAS: lost, location error – San Benito is in Baja | |
| Labridae | Halichoeres chierchiae# | Di Caporiacco, 1947 | H.nicholsi?# | Yes | Yes | Voucher, ID | No voucher; ID = H. nicholsi? |
| Labridae | Halichoeres dispilus# | (Günther, 1864) | H.insularis?# | Yes | Yes | Voucher, ID | No voucher; ID = H. insularis? |
| Labridae | Halichoeres melanotis# | (Gilbert, 1890) | H. sanchezi# | Yes | Yes | voucher, ID | SIO: not in catalog; ID = H. sanchezi (sister species) |
| Labridae | Halichoeres semicinctus | (Ayres, 1859) | H. insularis# | Doubt | – | ID, lost | LACM*; CAS: lost |
| Labridae | Thalassoma lutescens | (Lay & Bennett, 1839) | T.grammaticum | Yes | Reject | ID | T.lutescens in TEP = misidentified T. grammaticum |
| (Scarinae) | Nicholsina denticulata# | (Evermann & Radcliffe, 1917) | Calotomus carolinus | Yes | Yes | ID, voucher | USNM*; CMNFI*; CPUM*; CC10: no voucher |
| (Scarinae) | Scarus perrico | Jordan & Gilbert, 1882 | S.compressus? | Yes | Yes | voucher, location | SIO: not in catalog; MNHN: Gulf of California; no primary source |
| Labrisomidae | Labrisomus multiporosus | Hubbs, 1953 | L. socorroensis# | Yes | Yes | ID | UBC: in catalog as L. socorroensis |
| Labrisomidae | Labrisoums xanti# | Gill, 1869 | L.socorroensis# | Yes | Yes | ID | LACM*: L. socorroensis |
| Labrisomidae | Malacoctenus margaritae# | (Fowler, 1944) | Labrisomus? | Yes | Yes | ID, voucher | CC10; DMF; no voucher, = juvenile Labrisomus? |
| Lutjanidae | Lutjanus novemfasciatus | Gill, 1862 | – | – | ID | H20 video: unidentifiable; VL not mentioned | |
| Mobulidae | Mobula mobular | (Bonaterre, 1788) | Doubt | Yes | voucher | BG20: no evidence; synonym of M. mobular | |
| Mugilidae | Mugil curema# | Valenciennes, 1836 | M.setosus | Yes | Yes | synonym | Br19: TEP population is synonym; both species listed by Fourr & DMF |
| Mugilidae | Mugil thoburni# | (Jordan & Starks, 1896) | – | Yes | location | Albatross 1889: USNM -location Galápagos; a Galápagos endemic | |
| Mullidae | Pseudupeneus grandisquamis | (Gill, 1863) | Mulloidichthys dentatus | Yes | Yes | ID | SIO* |
| Muraenidae | Enchelynassa canina | (Quoy & Gaimard, 1824) | Gymnothorax sp. | – | – | ID | CPUM* |
| Muraenidae | Gymnothorax mordax | (Ayres, 1859) | G. flavimarginatus?# | – | – | ID, voucher | CB88: identification undecided; G. flavimarginatus & G. mordax coloration similar |
| Muraenidae | Gymnothorax undulatus | (Lacepède, 1803) | Yes | Yes | voucher | B&R: no mention; no primary source | |
| Muraenidae | Muraena argus | Yes | Yes | voucher | No primary source | ||
| Muraenidae | Muraena clepsydra | Yes | Yes | voucher | No primary source | ||
| Muraenidae | Uropterygius marmoratus | (Lacepede, 1803) | U.macrocephalus | – | – | ID, location | CMNFI*; U. marmoratus; Indo-west Pacific, not known |
| Muraenidae | Uropterygius polystictus | Myers & Wade, 1941 | Yes | Yes | voucher | CAB; no primary source | |
| Muraenidae | Uropterygius versutus | Bussing, 1991 | Yes | Yes | voucher | B91: no mention; no primary source | |
| Nematistiidae | Nematistius pectoralis | Gill, 1862 | Yes | Yes | location | MNHN: Gulf of California; no primary source | |
| Ophichthidae | Myrichthys tigrinus | Girard, 1859 | M. pantostigmius# | – | Yes | ID, location | SIO*; LACM*; M. tigrinus rename is M. xystrurus; MNHN: Gulf of California |
| Opistognathidae | Opistognathus punctatus | Peters, 1869 | Yes | Yes | location | SIO: outside study area | |
| Opistognathidae | Opistognathus rhomaleus | Jordan & Gilbert, 1881 | Yes | Yes | location | SIO: specimen = aquarium eggs, vague location | |
| Opistognathidae | Opistognathus rosenblatti | Allen & Robertson, 1991 | Yes | Yes | location | SIO: location suspect | |
| Oplegnathidae | Oplegnathus insignis | (Kner, 1867) | – | Reject | location | DMF: no source; Galápagos/Peru species | |
| Ostraciidae | Acanthostracion notacanthus | (Bleeker, 1863) | – | – | location, lost | CAS: lost; Atlantic species | |
| Polynemidae | Polydactylus approximans | (Lay & Bennett, 1839) | – | Reject | voucher | DMF: no source | |
| Polynemidae | Polydactylus opercularis | (Gill, 1863) | – | – | lost | CAS: lost | |
| Pomacanthidae | Centropyge bicolor | (Bloch, 1787) | Chaetodontoplus mesoleucus | – | – | location | Albatross 1889: USNM, Indo-west Pacific species, ledger says St. Lucia |
| Pomacentridae | Abudefduf concolor | (Gill, 1862) | Abudefduf declivifrons | Doubt | Reject | location, lost | CAS: lost; southern TEP species; A.declivifrons occurs in Mexico |
| Pomacentridae | Chromis limbaughi | Yes | Yes | voucher, lost | CAS: lost; no primary source | ||
| Pomacentridae | Chromis punctipinnis | (Cooper, 1863) | – | – | location | J&M: NHMUK location = San Benito (Baja), not San | |
| Pomacentridae | Hypsypops rubicundus | (Girard, 1854) | Yes | Yes | voucher, lost | CAS: lost; no primary source | |
| Sciaenidae | Cynoscion xanthulus | Jordan & Gilbert, 1882 | Yes | Reject | location | LACM: mainland site | |
| Sciaenidae | Roncador stearnsii | (Steindachner, 1875) | – | – | location | CAS: location = Baja | |
| Scorpaenidae | Scorpaena tierrae | Hildebrand, 1946 | S.mystes | – | – | ID, location | UBC*; Peruvian species. |
| Scorpaenidae | Scorpaena sp. | – | – | ID | LACM*: too small for reliable identification | ||
| Scorpaenidae | Scorpaenopsis gibbosa | (Bloch & Schneider, 1801) | – | – | location, lost | CAS: lost; Indian Ocean endemic | |
| Scorpaenidae | Sebastes carnatus | (Jordan & Gilbert, 1880) | – | Reject | location | Albatross 1889: USNM, temperate NE Pacific species, location unreliable | |
| Serranidae | Diplectrum euryplectrum | Jordan & Bollmann, 1890 | Serranidae | Yes | Yes | ID, lost, location | LACM*: too small for reliable identification; Albatross 1889: USNM, specimens lost, location unreliable |
| Serranidae | Serranus psittacinus# | Valenciennes, 1846 | S socorroensis# | – | Yes | ID | S&H: predates description of S.socorroensis |
| Sparidae | Calamus brachysomus | (Lockington, 1880) | – | – | voucher | no primary source | |
| Sphyraenidae | Sphyraena argentea | Girard, 1854 | Yes | Yes | lost, ID | SIO: lost, too small for reliable identification | |
| Syngnathidae | Syngnathus sp. | Bryx veleronis# | – | – | ID | LACM* | |
| Synodontidae | Synodus hoshinonus | Tanaka, 1917 | – | – | lost, location | CAS: lost; Indo-west Pacific species | |
| Tetraodontidae | Guentheridia formosa | (Günther, 1870) | Yes | Reject | location | Albatross 1889: USNM: location = Panama | |
| Triglidae | Prionotus albirostris | Jordan & Bollman, 1890 | Doubt | – | voucher | No primary source |
CPUM is the only Mexican museum or research collection from which we have examined RNP specimens to verify their identification and metadata. Fricke et al. (2024) provide comprehensive information on species occurrences in Mexican ichthyological collections in their checklist of Mexican fishes. They list Mexican collections for some of the reef-associated species in the Castro- Aguirre and Balart (2002) checklist that we included in Table 2 (and Table 3) here, but do not indicate whether those collections contain specimens from the RNP of those species: Carcharhinus brachyurus, Carcharhinus leucas, Aetobatus laticeps, Mobula mobular, Muraena argus, Uropterygius polystictus, Serranus psittacinus, Gnathanodon speciosus, Nematistius pectoralis, Abudefduf declivifrons, Halichoeres dispilus, and Labrisomus multiporosus. If any of those collections include RNP specimens those should be examined to verify identities and collection metadata. Species accounts of selected unaccepted species Below are species accounts of 18 unaccepted species that expand on the reasons for their rejection summarized in Table 2.
| Family | Problem | Scientific name | Name authority | Fourriére et al., 2016 | Del Moral-Flores et al., 2016 | Sources of information about problems |
|---|---|---|---|---|---|---|
| Balistidae | Voucher | Pseudobalistes naufragium# | (Jordan & Starks, 1895) | Yes | Yes | CC10: observation – possible misidentification |
| Belonidae | ID, lost, source | Strongylura exilis# | (Girard, 1854) | Yes | Yes | CB88: possible misidentification; CMNFI*: misidentified Platybelone pterura; CAS: lost; SIO: unlocatable |
| Carangidae | eDNA only | Decapterus macrosoma | Bleeker, 1851 | eDNA: 14r CO1 | ||
| Carapidae | ID, location | Echiodon exsilium# | Petit, 1934 | SIO: misidentified as Carapus mourlani, location suspect | ||
| Chanidae | eDNA only | Chanos chanos | (Fabricius, 1775) | eDNA: 21r CO1; MNHN is Gulf of California | ||
| Congridae | location | Gorgasia punctata# | Meek & Hildebrand, 1923 | SIO: location suspect | ||
| Opistognathidae | location | Opistognathus fossoris# | Bussing & Lavenberg, 2003 | SIO: location suspect | ||
| Paralichthyidae | eDNA only | Paralichthys californicus | (Ayres, 1859) | eDNA: 21r CO1, 735r 16S | ||
| Scombridae | ID | Sarda chilensis# | (Cuvier, 1832) | Yes | Yes | UBC: not in UBC online database |
Acanthurus achilles. The only record of this species in the RNP is an unpublished grey literature report by Chan (1974), which was accepted by Del Moral-Flores et al. (2016) but not mentioned by Fourriére et al. (2016). That account demonstrated an obvious lack of familiarity with RNP fishes as most species were only identified to genus. It described A. achilles as common at San Benedicto and in Hawaii but did not mention Acanthurus nigricans, which is common at both the RNP and the Hawaiian Is. Other records in that article likely represent confusion of names, including multiple records of E. analogus (as grouper) but no mention of the common E. clippertonensis (as E. labriformis since this publication preceded the description of E. clippertonensis) and of F. longirostris (a Hawaiian species) but not F. flavissimus. Hence, we attribute this report of A. achilles to misidentification and regard it as erroneous. Although there are a few validated records of A. achilles from Clipperton and Baja California, there is no evidence of a resident population in the TEP.
Carcharhinus leucas. The only record of this species in the RNP is a single juvenile Carcharhinus collected in 1982 by Castañeda-Beltrán (1988) at Clarion and recorded as C. leucas. Becerril-García et al. (2020) cited Castañeda- Beltrán as doubtful and do not include C. leucas in their curated inventory of RNP elasmobranchs. Juveniles of morphologically similar Carcharhinus species can be hard to identify, the very short description of the individual collected by Castañeda-Beltrán (1988) is not definitive for C. leucas and could represent any of several other Carcharhinus species known from the RNP. Hence, this record of C. leucas at the Revillagigedo Islands is not accepted.
Carcharhinus melanopterus. The specimen collected in Socorro in 1959 that was deposited in the UBC collection under this name has been destroyed. However, there are confirmed records of this species at Cocos Is. (Fourriére et al., 2017) so it is possible that this record in the RNP is correct. We record its presence in the RNP as unaccepted as there is no way to resolve whether or not the UBC specimen was correctly identified.
Gobiesox aethus. Briggs (1951) described G. aethus based on a single specimen from Clarion and to date it, along with G. canidens, has been regarded as an endemic (Del Moral-Flores et al., 2016; Fourriére et al., 2016). Torres-Hernández et al. (2022) assessed DNA barcode sequences of 6 specimens from Clarion and 19 from Socorro as part of a phylogeographic study of the widely distributed Gobiesox adustus. Sampling of additional individuals has brought the total number of CO1 sequences from the RNP to 45 and indicates that only 1 species, an endemic, is present there. Since the description of G. canidens preceded that of G. aethus by Briggs (1951), the likely name of the RNP endemic population is G. canidens and G. aethus is a synonym.
Halichoeres chierchiae, Halichoeres dispilus, and Halichoeres melanotis. Juveniles of these 3 species, which are very common inhabitants on mainland reefs, resemble, respectively, juveniles of the Revillagigedos form of H. nicholsi and the Revillagigedo endemics H. insularis and H. sanchezi. Given that the records of all 3 in the RNP are based entirely on unvouchered observations of each species, misidentification is a reasonable explanation for all 3 cases and they likely represent invalid records.
Kyphosus lutescens. This species was named by Jordan and Gilbert (1892) based on a single, all-yellow specimen collected at Socorro. Recently it was synonymized with K. sectatrix, a pantropical species, by Knudsen and Clements (2013, 2016) based on morphological comparisons. However, Del Moral-Flores et al. (2016) included K. lutescens as an RNP endemic and rejected K. sectatrix in their inventory. In contrast, Fourriére et al. (2016) included only K. sectatrix in their inventory. Coloration of RNP specimens varies from all-yellow to all white to all- grey, with many fish showing intermediate patterns with varying degrees of 2 or 3 colors on different parts of the body (Supplementary material: Files S1-S4; iNaturalist images; Robertson & Allen, 2024; Valencia-Méndez et al., 2018). Similar local color variation is known in other populations of K. sectatrix elsewhere in the Indo-Pacific. In addition, newly available DNA barcodes of yellow and grey morphs from the RNP by ODD (unpublished data) show that they belong to a single haplogroup and support Knudsen and Clement’s (2016) conclusion that K. lutescens is a synonym of K. sectatrix. In the eastern Pacific K sectatrix occurs not only at the Revillagigedos, but also Baja California, the Galápagos, Isla del Coco, Roca Alijos and Guadalupe Island.
Labrisomus xanti. See comments about L. socorroensis in endemics section.
Malacoctenus margaritae. The only record of this labrisomid species is from Chávez-Comparan (2010), who did not mention observing any other species of labrisomids. Del Moral-Flores et al. (2016) pointed out that this species is a southern TEP species occurring on the mainland from Costa Rica southwards and proposed that what Chávez-Comparan (2010) saw could have been Malacoctenus mexicanus, which occurs on the Mexican mainland. Fourriére et al. (2016) also referred to this record as M. mexicanus. We suggest that more likely it was either a juvenile Labrisomus or even E. exsul. In the absence of vouchers confirming otherwise we exclude M. margaritae/mexicanus from the accepted species set.
Mugil curema. See M. setosus in accepted section.
Mugil thoburni. This species was described in 1896, from specimens collected in the Galápagos and currently is regarded as a Galápagos endemic (Victor, Grove et al., 2024). Both Del Moral-Flores et al. (2016) and Fourriére et al. (2016) recorded it in the RNP, based on a USNM record that is from a collection with many erroneous location records (see above) and thus is not considered a validated record.
Nicholsina denticulate. The few records of this species in the RNP consist of specimens deposited in CMNFI (2 individuals collected in 1968) and USNM (3 individuals collected in 1994), plus observations by Chávez-Comparan (2010) and 4 possible individuals collected at Socorro by ODD. Reexamination of those 2 sets of museum specimens, supported by DNA results of the 4 ODD fish (unpublished data), show that they are all Calotomus carolinus. Thus, the only remaining potential records of N. denticulata are from observations, and Chávez- Comparan (2010) listed neither species as common. Young individuals of both species have very similar forms and mottled brownish camouflage coloration and can easily be confused, while IP individuals are also quite similar in form and color. Only the TP males of those 2 species are readily distinguishable by color patterns (Robertson & Allen, 2024). Given the fact that the museum specimens were incorrectly identified we think that misidentification is also likely with those observational records. We only observed and photographed C. carolinus at the RNP in 2022 and the only photographs in iNaturalist are of C. carolinus (n = 37 taken between 2017 and 2025). Hence, we list N. denticulata as part of the non-accepted group.
Serranus psittacinus. See S. socorroensis in endemics section.
Sufflamen fraenatum. Del Moral-Flores et al. (2016) suggested that specimens of S. fraenatum from the TEP represent misidentified Sufflamen verres, which is common throughout the region and the RNP. Specimens named S. fraenatum (Balistes capistratus and Pachynathus capistratus are synonyms) were collected at Clarion and Socorro in 1897 by Jordan and McGregor (1898). However, in which museum Jordan and McGregor’s specimens are housed is not clear. The USNM has specimens of B. capistratus renamed as S. verres, that were collected at Socorro and Clarion in by the Albatross expedition of 1883-1889, i.e., before S. verres was described in 1904. GBIF also lists 5 museum records from the Gulf of California and the Galápagos, but only the latter postdates the description of S. verres. Reexamination of specimen data for a 1957 Seckenburg Museum “S. fraenatum” from the Galápagos showed that it had already been reidentified as S. verres. We agree with Del Moral-Flores et al. (2016) that the S. fraenatum records from the RNP are misidentified S. verres. Although Palacios-Morales et al. (2014) identified a juvenile Sufflamen collected on the southern coast of Mexico as S. fraenatum, the color pattern of that individual resembles that of juvenile S. verres rather than juvenile S. fraenaum and genetic reassessment (by ODD, unpublished data) shows that it is indeed an S. verres. Hence, there is no convincing evidence of any occurrences of S. fraenatum in the TEP. However, that species is distributed throughout the Indo-Central Pacific, including isolated islands and archipelagos, indicating a strong capability for dispersal. It occurs at all major island groups on the western edge of the East Pacific Barrier, including the Line Islands, which is the most likely location to dispatch propagules of transpacific migrants towards the TEP on the equatorial countercurrent. Hence, it does have the geographic and biological potential to be a transpacific migrant to the TEP and might well be recorded there in the future.
Tylosurus pacificus. Del Moral-Flores et al. (2016) and Fourriére et al. (2016) recorded T. melanotus and T. pacificus in the RNP. The T. pacificus records are based on misidentified T. melanotus (see account of T. melanotus among accepted species).
Xanthichthys lineopunctatus and Xanthichthys auromarginatus. Xanthichthys mento was described from a specimen from Clarion by Jordan and Gilbert (1882), who made no mention of X. lineopunctatus, which has a generally similar color pattern to X. mento. Records labelled as X. lineopunctatus from the RNP, collected in the mid-20th century, still exist in several museums. Our reexaminations of such specimens from SIO, MCZ and CUMV now shows that they are X. mento. In any case, X. lineopunctatus is an Indo-west Pacific species, with its nearest population 11,000 + km from the Americas (Matsuura, 2022), so is highly unlikely to occur in the TEP. However, 3 other species of Xanthichthys: X. auromarginatus, Xanthichthys caeruleolineatus and Xanthichthys greenei, do occur at the Line Islands on the western edge of the East Pacific Barrier and X. caeruleolineatus occurs as a vagrant at Cocos and Galápagos. Giddens et al. (2019) recorded deep-living fishes using dropcams in the RNP and other offshore islands in the TEP. Inspection by DRR of a relatively low-quality image from a video that they identified as X. auromarginatus that was kindly provided by J. Giddens indicates it was a S. verres and not a Xanthichthys species.
Review of an RNP specimen labelled Xanthichthys ringens, an Atlantic species, in CMNFI, also showed that it is X. mento. The only species we have observed or collected in the RNP between 2012-2025 is X. mento, which is very common at all 4 islands, and there are numerous records of this species in the RNP from the 1940s to the present, including many iNaturalist images. Based on this information we conclude that the only species in the RNP, and elsewhere in the TEP, are X. mento as a resident, and X. caeruleolineatus as a vagrant at southern sites in the TEP.
Species whose presence in the RNP requires confirmation
In addition, there are 9 potential occurrences in need of confirmation through more robust voucher data (see Table 3).
Pseudobalistes naufragium. The only record of this species in the RNP is an observation by Chávez-Comparan et al. (2010), who did not list it among the common species they observed. They also did not record the life stage of the individual(s) observed. While adults of P. naufragium usually are readily distinguishable from adults of Balistes polylepis, another plain colored balistid that is common and found throughout the RNP, the juveniles of those 2 species can appear quite similar. Hence, without a voucher, the record of P. naufragium is listed as unresolved.
Strongylura exilis. There are several records of this species in the RNP. Three needlefishes identified as S. exilis in the UBC catalog that were collected both during and shortly after the 1954-1958 activity discussed by Ricker (1959) that subsequently were transferred to CMNFI were reexamined and identified as Platybelone pterura. There is a 1925 CAS record from Clarion, but the specimen has been lost and there is no other information about it in the CAS online catalog. A specimen relating to the 1946 record in the SIO database, of a synonym, Tylosurus stolzmanni, cannot be located and the record has been removed from the catalog. Castañeda-Beltrán (1988) collected 2 needlefishes from Clarion in 1982 that he named S. exilis and recorded that they ranged between 83 to 100 cm TL. Although S. exilis is known to reach a maximum size of 91cm TL (Love et al., 2021), 2 other belonids confirmed from the RNP that reach 100 cm are T.crocodilus and T. melanotus, either of which might have been named as S. exilis by Castañeda-Beltrán (1988). It should also be noted that there has been confusion about the application of correct genus names to TEP Tylosurus and Strongylura species collected during the 20th century that are in various museum collections: e.g., Tylosurus exilis (CAS), Strongylura fodiator (UBC), T. stolzmanni (SIO), and Strongylura galapagensis (ANSP), the last a synonym of T. pacificus. Given the absence of confirmed records, the confusion about identifications and names of needlefish specimens and a lack of putative records since 1982, we consider the occurrence of S. exilis in the RNP to be unresolved.
Echiodon exsilium, Gorgasia punctata and Opistognathus fossoris. The records of these 3 species, which are known to occur in the Gulf of California and the mainland Mexican coast, come from a single small lot of 4 species obtained by A. Kerstitch (University of Arizona) that, according to the SIO catalog, was collected at Clarion on June 1, 1993. That specimen lot, initially in the University of Arizona collection, was later moved to SIO. When they were reexamined by author BF the identities of G. punctata and O. fossoris were confirmed but the E. exsilium was found to have been misidentified as Carapus mourlani. The only other species collected by Kerstitch during 1993 listed in the SIO catalog is a different lot of 2 Pontinus vaughani, also collected on June 1, but, according to the SIO catalog, “doubtfully obtained at La Paz”, southern Baja. If Kerstitch collected fish at the Revillagigedo islands in 1993, it seems likely that there would be more than 1 small lot containing 3 species for which there are no other records in the RNP. Given this conflicting information about the June 1 collecting location and the lack of information about other fish specimens collected by Kerstitch from currently unknown locations during 1993, we regard these 3 sole- source records as requiring confirmation.
Sarda chilensis. Del Moral-Flores cited Ricker (1959) as the source of the only record of this species in the RNP. Ricker (1959) indicated that a specimen collected at Socorro was deposited in the UBC collection. However, the UBC online database in GBIF (UBC database, GBIF) does not contain any records of this species from the RNP.
Fricke et al. (2024) list Mexican collections for 4 of the species included in Table 3: P. naufragium, S. exilis, E. exsilium, and S. chilensis. If any of those collections include RNP specimens of those species, they should be examined to verify identities and collection metadata.
Using the UNESCO eDNA study we added 3 species not previously recorded from the RNP to the “requiring confirmation” group, due to the potential for eDNA to be derived from larvae rather than adults and to relatively low numbers of reads of ASVs of CO1 for each of those species: Decapterus macrosoma (the 2 other members of the genus found in the TEP were also detected by eDNA from the RNP, as well as by other voucher information), Chanos chanos (a monospecific family), and Paralichthys californicus (whose GenBank reference sequence we determined is correct). Decapterus macrosoma is known from the Galápagos and C. chanos from all the other TEP oceanic islands. Paralichthys californicus occurs on the Pacific coast of Baja as far south as Magdalena Bay and its larvae could be occasionally swept to the RNP by the southerly flow of the California Current.
Endemism of the RNP reef-associated fish fauna
Fourriére et al. (2016) listed 12 species as RNP endemics, while Del Moral-Flores et al. (2016) listed 23. Information presented in the species accounts below indicate that while 12 of the species mentioned by both 2016 inventories are validate endemics, 4 more mentioned only by Del Moral-Flores et al. (2016) are also endemics, and another 7 species, one of them included on both inventories and another only on the Del Moral-Flores et al. (2016) inventory, are probable endemics. In addition, another 3 species (one of them on the Del Moral-Flores et al. (2016) inventory as an endemic) are potential endemics (Table 4). However, we disagree with the listing of 7 other species as endemics in one or both 2016 inventories, as explained in the species accounts of unaccepted endemics, below. In combination our investigations indicate that there are 16 described and named endemics, another 7 probable endemics, only 2 of them named, and 4 potential endemics (see Table 4 and species accounts below). Those included some species that have been recorded from the mainland but evidently lack self-sustaining populations there.
| Family | Species | Endemic Status | Known from | Fourriére et al. (2016) | Del Moral-Flores et al. (2016) |
|---|---|---|---|---|---|
| Blenniidae | Hypsoblennius proteus | Known | B, C, P, S | Yes | Yes |
| Chaenopsidae | Acanthemblemaria mangognatha | Known | B, C, S | Yes | Yes |
| Dactyloscopidae | Dactyloscopus insulatus | Known | B, C, S | Yes | Yes |
| Gobiesocidae | Gobiesox canidens | Known | C, P, S, (B?) | Yes | Yes |
| Gobiesocidae | Tomicodon absitus | Known | B, P, S | Yes | Yes |
| Gobiesocidae | Tomicodon sp. | Probable | C | New data | New data |
| Gobiidae | Bathygobius ramosus longipinnis | Probable | C, S, (B?) | No | Yes |
| Gobiidae | Chriolepis sp. | Probable | B, C, S | New data | New data |
| Gobiidae | Coryphopterus urospilus | Potential | B, C, S | New data | New data |
| Gobiidae | Lythrypnus cf. dalli | Probable | C | New data | New data |
| Gobiidae | Lythrypnus insularis | Known | B, C, S | Yes | Yes |
| Haemulidae | Anisotremus perezponcedeleoni | Known | B, C, P, S | New data | New data |
| Labridae | Halichoeres insularis | Known | B, C, P, S | No | Yes |
| Labridae | Halichoeres nicholsi | Probable | B, C, S | New data | New data |
| Labridae | Halichoeres sanchezi | Known | B, S | New data | New data |
| Labridae | Xyrichtys sp. | Probable | C, S, (B & P?) | Yes | Yes |
| Labrisomidae | Labrisomus socorroensis | Known | B, C, S | Yes | Yes |
| Pomacanthidae | Holacanthus clarionensis | Known | B, C, P, S | No | Yes |
| Pomacentridae | Stegastes redemptus | Known | B, C, S | No | Yes |
| Sciaenidae | Pareques sp. | Probable | B, C, S | New data | New data |
| Serranidae | Rypticus courtenayi | Known | B, C, P, S | Yes | Yes |
| Serranidae | Serranus socorroensis | Known | B, C, P, S | Yes | Yes |
| Soleidae | Aseraggodes herrei | Potential | B, C, S | No | No |
| Syngnathidae | Bryx clarionensis | Potential | C, S, (B?) | No | Yes |
| Trypterigiidae | Axoclinus multicinctus | Known | B, C, S | Yes | Yes |
| Trypterigiidae | Enneanectes exsul | Known | B, C, S | Yes | Yes |
Named endemic species accounts
Acanthemblemaria mangognatha. There are no records of this species, described by Hastings and Robertson (1999) as a Revillagigedo endemic, from anywhere in the TEP other than the RNP. Museum specimen records of 3 mainland congeners, Acanthemblemaria crockeri, Acanthemblemaria hancocki, and Acanthemblemaria macrospilus, at the RNP all predate that description. None of them were accepted by Del Moral-Flores et al. (2016) or Fourriére et al. (2016) and specimens of each have been shown to be A. mangognatha (Hastings & Robertson, 1999; our own specimen reviews).
Anisotremus perezponcedeleoni. The RNP population previously known as Anisotremus interruptus was described as an RNP endemic, A. perezponcedeleoni, by Acevedo-Álvarez et al. (2021), who used both morphological and genetic data to arrive at that conclusion.
Axoclinus multicinctus and Enneanectes exsul. These are 2 small, cryptobenthic triplefins. A. multicinctus was described by Allen and Robertson (1992c) and E. exsul by Rosenblatt et al. (2013). Those are the only 2 tripterygiids currently known from the RNP and recorded there since 2000. Records from the 1950s and 1960s of Axoclinus carminalis, Axoclinus lucillae, Axoclinus sp., Enneanectes sexmaculatus (= A. carminalis) and Enneapterygius sp. (a genus that does not occur in the neotropics) have not been curated since being deposited under those names in UBC and most likely are of either or both A. multicinctus and E. exsul as there are no records of any other tripterygiids being collected in the RNP since those UBC collections. The same applies to the 1925 CAS collection of “Tripterygion dubius” in the RNP, a species name that does not appear in Eschmeyer’s Catalog of Fishes (Fricke et al., 2026) from a genus that does not occur in the Americas.
Dactyloscopus insulatus. This Revillagigedos endemic species, which is known from all 3 main islands in the RNP, was described as an endemic subspecies of Dactyloscopus pectoralis by Dawson (1975) and subsequently raised to the species level by Hastings and Springer (2009). There are no records of this species from anywhere in the TEP other than the RNP and no other members of the genus have recorded from the RNP. We obtained images and specimens at Socorro in 2022.
Gobiesox canidens. Briggs (1951) described 2 species of Gobiesox from the RNP: G. aethus, based on a single specimen from Clarion, and G. canidens, based on 5 specimens from Socorro. To date both species have been regarded as endemics (Del Moral-Flores et al., 2016; Fourriére et al., 2016; Fricke et al., 2024, 2026). Torres- Hernández et al. (2022) examined genetic relationships among populations of G. adustus (Jordan & Gilbert, 1882), assessing DNA barcode sequences of 155 individuals across the great majority of its geographic range, from Baja California to Ecuador and Isla del Coco, and including 6 specimens from Clarion and 19 from Socorro. The Revillagigedo haplotypes formed a single haplogroup segregated by a p-distance of 2.6% from the Baja California haplogroup, the nearest other haplogroup, and there are no mainland haplotypes in the RNP. Subsequent sampling of additional individuals has brought the total number of sequences from the RNP to 45 and reinforces earlier indications that only 1 haplogroup and species is present there. A more comprehensive, integrative taxonomic analysis of the Revillagigedo population will be presented by ETH at a later stage. Gobiesox canidens is the likely name of the RNP endemic population since its description preceded that of G. aethus in Briggs (1951). Images show that fish from the RNP resemble G. adustus in coloration, a dark brown fish with thin pale blue-white irregular lines crossing the head and running longitudinally along the body (Supplementary material: Files S1, S4, and G. adustus color)
Halichoeres insularis. This species was described by Allen and Robertson (1992a) from specimens collected at Socorro in 1991. It is now known to also occur at Clarion and San Benedicto. While Del Moral-Flores et al. (2016) accepted this species as an endemic, Fourriére et al. (2016) did not, without explanation. This was likely due to supposed records from Alijos Rocks and Guadalupe Island indicated by Robertson and Allen (2015). Those were predicated on the assumption that a photo of a terminal phase male taken at Guadalupe (Guadalupe Halichoeres) was in fact H. insularis. However, during the 2022 expedition DRR searched for but was unable to find any TPs like that from Guadalupe, despite the superabundance of H. insularis in the RNP (aggregations of scores to hundreds of individuals were common at all 3 major islands). The only color phases seen in 2022 were those described by Allen and Robertson (1992a). Hence, we conclude that there is another, as yet undescribed species of Halichoeres related to H. insularis at Guadalupe and perhaps Alijos Rocks. Based on that we treat H. insularis as an RNP endemic. The taxonomic status of the Guadalupe population remains to be investigated.
Halichoeres sanchezi. This distinctively colored species was recently described as a new endemic wrasse from the RNP, currently known only from San Benedicto and Socorro (Victor, Frable et al., 2024) and nowhere else in the TEP. Halichoeres melanotis is the sister of H. sanchezi and reports of the former in the RNP made before 2024 most likely actually refer to H. sanchezi.
Holacanthus clarionensis. This species was listed as an endemic by Del Moral-Flores et al. (2016), but not by Fourriére et al. (2016). It is common at all 4 Revillagigedo islands, where photographs of aggregations of scores of individuals are easily obtained. Records on the mainland are scattered from southern Mexico to the central Gulf of California, along southwestern Baja up to Cedros Island and at Guadalupe Island. As of December 2024, GBIF had 26 primary-source records from the mainland, from 1959 to the present; those include 12 photographs from iNaturalist. AME&CJE photographed single fish at Cabo Pulmo and Cabo San Lucas in 2021-2022. All photographic records from sites other than the RNP are of single isolated adults. The similarly sized sister of H. clarionensis, H. passer, can live up to 20 years (Fernández-Rivera et al., 2016) and mainland records of H. clarionensis that have accumulated over the past 85 y most likely refer to single, long-lived isolated waifs rather than members of a self- maintaining mainland population. Hence, we regard H. clarionensis as a Revillagigedos endemic. We include an image of a likely hybrid of H. passer X H. clarionensis (Supplementary material: File S7) collected at Socorro during the 2022 expedition. Coauthor AAB obtained an image (video screengrab) of another such hybrid at Socorro in 2024.
Hypsoblennius proteus. This species was described by Krejsa (1960) as endemic to the RNP, based on 37 specimens in 4 museums from all 4 RNP islands. He compared those to 217 specimens of Hypsoblennius brevipinnis from sites scattered throughout Baja California and the Gulf of California, Cocos, Panama, Colombia, Ecuador, Peru and the Galápagos. He did not mention finding H. proteus anywhere except the RNP or finding H. brevipinnis in the RNP. As noted by both 2016 inventories, H. proteus is endemic to the RNP and the only member of its genus known from there.
Labrisomus socorroensis. This species was described as an RNP endemic by Hubbs (1953), based on 2 specimens collected at Socorro. Hubbs noted its similarity to L. xanti, including in the structure of its (preserved) color pattern and distinguished the 2 based on patterns of scalation on the side of the head. Hubbs (1953) also examined 559 L. xanti and described L. multiporosus, based on 253 specimens, without mentioning the occurrence of either species in the RNP. A LACM specimen from Clarion collected in 1971 that was listed as L. xanti was reexamined and reidentified as L. socorroensis. As of January 2026, there are 9 images of live Labrisomus from Socorro in iNaturalist, 7 labelled L. xanti and 2 labelled L. socorroensis. The Ricker (1959) specimen recorded as L. multiporosus with a UBC number is listed in the UBC catalog as L. socorroensis, and all of that museum’s specimens of Labrisomus from the RNP are listed as L. socorroensis (see GBIF).
ODD (unpublished data) DNA barcoded 28 individuals of Labrisomus from the RNP; 25 of those belong to a haplogroup composed solely of RNP fish, separated from a haplogroup containing mainland L. xanti by a p-distance of 0.6%. In addition, 2 RNP specimens have haplotypes found in the L. xanti haplogroup. There are various scenarios that could explain the existence of a few “L. xanti” haplotypes in the RNP, including its presence there as vagrants, hybridization with its close relative, L. socorroensis, and an early stage of speciation by L. socorroensis (see Craig et al., 2006 for a detailed discussion of a similar situation in Epinephelus clippertonensis). Until this situation is examined in greater depth, involving morphology and other genes, we suggest that L. socorroensis should be regarded as endemic to the RNP and the presence of L. xanti as unresolved.
Lythrypnus insularis. Both Del Moral-Flores et al. (2016) and Fourriére et al. (2016) listed this species among the RNP endemics. Bussing (1990) described this species from 54 specimens collected at Clarion and 33 collected at Socorro, and this is the only member of the genus he noted as occurring in the RNP. RNP records of congeners (Lythrypnus sp., Lythrypnus pulchellus, Lythrypnus rhizophora, and Lythrypnus zebra) in the collections of LACM, SIO and CAS from the RNP all refer to fish collected in 1925, 1953, 1955, 1959, 1970 and 1971 and named as such long before Bussing’s description. A few records from 1971 were reidentified as L. insularis after Bussing’s description of that species. Lythrypnus insularis has not been recorded anywhere other than at the RNP. In 2022 we collected 7 L. insularis from Socorro, 27 from Clarion and 1 from San Benedicto. DNA barcodes of 7 L. insularis from Socorro and 15 from Clarion collected in 2022 by ODD (unpublished data) show only 1 species present, with a suggestion of some separation of the Socorro and Clarion populations. Del Moral-Flores et al. (2016) and Fourriére et al. (2016) both recorded L. pulchellus and L. zebra as present in the RNP. However, given the lack of records of either of those species since 1971 when the L. insularis specimens described by Bussing (1990) were collected, we conclude that L. insularis is the only named species among that group that is present in the RNP. A limited number of DNA barcodes of L. pulchellus, L. rhizophora, and L. zebra as well as a large sample of L. insularis from Socorro and Clarion indicate distinct separation of the RNP populations, and that differences between L. insularis and the 3 others vary from 7.5% for L. rhizophora to 10.3% for L. pulchellus and 15.3% for L. zebra. The 2022 specimens and photographs of 50 + different individuals show the same basic, although variable, color pattern: body reddish anteriorly shading to dark grey posteriorly with 13-14 long thin, dark-edged iridescent vertical blue bars; head reddish to golden with varying degrees of red to gold spotting on side and lower part of head (lower spots with dark centers in some fish), with thin, dark-edge blue cross bars on nape and cheeks; iris reddish to gold; fins vary from transparent to grey, the unpaired fins darkest.
Rypticus courtenayi. This species was described as an RNP endemic by McCarthy (1979), in a paper reviewing the taxonomic status of the 3 TEP species: Rypticus bicolor, R. courtenayi, and Rypticus nigripinnis. McCarthy based his description of R. courtenayi on 80 specimens he examined from Clarion, San Benedicto and Socorro. He also examined 237 specimens of R. bicolor and 203 of R. nigripinnis and made no mention of either species at the RNP. Guimares (1999), in his review of the genus, also considered R. courtenayi to be the only species in the RNP. Major distinguishing features of the 3 species include the number of dorsal-fin spines, patterns of pores on the head and jaws, shape of the chin and color pattern. Five museum collections (LACM, USNM, SIO, CMNFC and CAS) from the RNP that predate McCarthy’s paper have specimens labelled as either R. bicolor (or its synomym Rypticus xanti) or R. nigripinnis. Only 1 (adult) among those specimens from LACM, CMNFI and SIO that were in an identifiable state has morphology consistent with R. nigripinnis rather than R. courtenayi. Prior to 2022 ODD obtained 12 specimens from the RNP, plus specimens of R. bicolor (n = 36) and R. nigripinnis (n = 7) from the mainland, Cocos, Galápagos and Clipperton. DNA barcodes (ODD unpublished data) clearly separate those 3 species into 3 haplogroups, with p-distances = 1.7% separating R. courtenayi from R. bicolor and 4.9% separating R. courtenayi from R. nigripinnis. During the 2022 expedition 2 juveniles were collected from Roca Partida. Although head pore patterns of those 2 were of the R. nigripinnis type their DNA barcodes clustered with those of R. courtenayi, indicating that differences in pore patterns described by McCarthy (1979) are not definitive for juveniles. The haplotype of 1 individual in a sample of 14 Rypticus from Clipperton also clustered with the R. courtenayi haplogroup, the remainder with R. bicolor. This was the only non-RNP R. courtenayi to be recorded either in museum collections or identified by DNA barcodes outside the RNP. The combination of morphology of McCarthy’s large sample of RNP fish and these DNA results strongly indicate that R. courtenayi is an RNP endemic, with occasional waifs reaching Clipperton Island and occasional waifs of R. nigripinnis reaching the Revillagigedo islands.
Serranus socorroensis. There are no records of S. socorroensis which was described from 5 Socorro specimens by Allen and Robertson (1992b) anywhere in the TEP other than the RNP. There are records of 2 other Serranus species from the RNP: S. psittacinus (as Prionodes fasciatus) by Snodgrass and Heller (1905), who simply mentioned it as occurring in those islands, without further comment or reference to a specimen; and a USNM record (USNM 94027) of S. aequidens from Socorro, which was cited by Fourriére et al. (2016). Our reexamination of that USNM specimen confirms that identification. Castro-Aguirre and Balart (2002) included S. psittacinus in the Revillagigedos fauna, citing Snodgrass and Heller (1905), as did Del Moral-Flores et al. (2016). Since there are no voucher specimens or other records of S. psittacinus from the Revillagigedos, and the structure of its color pattern resembles that of S. socorroensis, which was described long after Snodgrass and Heller’s paper, Heller and Snodgrass’s record most likely refers to S. socorroensis. We conclude that, at present, S. socorroensis and S. aequidens represent the only 2 confirmed species of Serranus known from the RNP islands, the former as an endemic.
Stegastes redemptus. This species is known from all Revillagigedo islands except Roca Partida. There are a few mainland records from southern Baja California: 5 in GBIF between 1960-2018, a photograph of a juvenile at Baja California taken in 2018 on iNaturalist, and photographs of a single individual taken at Ventana by CJE & AME in 2021. Hence, we conclude that the few mainland records are of single, isolated waifs, that there is no established population anywhere other than the RNP and that it is endemic to that park.
Tomicodon absitus. Briggs (1955) described T. absitus from 11 specimens collected at Socorro. This was considered to be the sole endemic member of the genus in the RNP by Del Moral-Flores et al. (2016), Fourriére et al. (2016), and Fricke et al. (2026). DNA barcoding of Revillagigedo specimens of the genus by ODD (unpublished data) shows that there are 2 haplogroups separated by a p-distance of 19.5%, 1 at Socorro (n = 22) and the other at Clarion (n = 52). Three other Tomicodon species have been recorded at the RNP at various times: Tomicodon eos, Tomicodon petersii, and Tomicodon zebra. The sequences of the 2 Revillagigedos lineages do not cluster with the sequences of the latter 2 of those 3 (Torres- Hernández et al., 2022), or with sequences of 3 other species of Tomicodon that occur in mainland Mexico but have not been recorded at the RNP, Tomicodon boehlkei, Tomicodon humeralis, and Tomicodon myersi. There are no sequences available for T. eos for comparison. The only possible validated record of those 6 non-endemic species in the Revillagigedos is of T. petersii, an LACM specimen from Socorro identified by Briggs (1955). These results indicate that there are 2 endemic Tomicodon species at the Revillagigedos, T. absitus at Socorro (and probably San Benedicto and Roca Partida) and Tomicodon sp., an undescribed species, from Clarion and that waifs of T. petersii may occasionally manage to recruit to the RNP. A more comprehensive, integrative taxonomic analysis of the 2 Revillagigedo populations will be presented by ETH at a later stage.
Probable endemic species accounts
Bathygobius ramosus longipinnis. This cryptobenthic species, which lives in intertidal habitat, was described as a subspecies endemic to the RNP by Ginsburg (1947). Del Moral-Flores et al. (2016) included it as an endemic subspecies, while Fourriére et al. (2016) referred to it simply as B. ramosus, which occurs on the mainland from the Gulf of California to Peru. However, Miller and Stefani (2001), which was cited in neither 2016 inventory, regarded it as a valid subspecies, based on morphological comparisons, but also noted similarities with the population at the Tres Marías Islands, in the mouth of the Gulf of California. ODD collected 15 individuals at both Clarion and Socorro, as well as substantial numbers of individuals from mainland sites throughout the rest of B. ramosus range, between Baja California and Ecuador. Together the Clarion and Socorro populations represent a very well-defined CO1 haplogroup (ODD unpublished data) that is well separated (by 3.6%) from B. ramosus haplogroups on the mainland. This indicates that the RNP population, with no other haplogroups present there, most likely is endemic. Although information is lacking on the genetic relationship of the Tres Marías population to those in the RNP and the mainland it seems unlikely that it will be closely allied to the RNP population, since the Tres Marías are much closer to the mainland (< 90 km) than to the RNP (500 km). The Galápagos and Clipperton are populated by other members of the genus (Miller & Stefani, 2001), while B. ramosus also occurs at Isla del Coco (Fourriére et al., 2017: ODD unpublished CO1 data).
Chriolepis sp. Neither 2016 inventory mentions any Chriolepis from the RNP. This undescribed cryptobenthic species was collected at Clarion and San Benedicto by the 2022 expedition and at San Benedicto in 2023 by ODD. There is an SIO collection of 39 specimens of Chriolepis from Clarion in 1955, by R. Rosenblatt et al., who labeled it Chriolepis n. sp. There also are LACM records of a Chriolepis sp. collected in 1939 (n = 8) and 1971 (n = 6) from Clarion and Socorro, but those specimens are listed in the catalog as lost, so are not included here. Findley (1983) examined the LACM specimens and concluded that those from the RNP likely represent an endemic but did not describe and name it because the specimens were so small. No other members of the genus have been recorded from the RNP. A description of this probable endemic species is in preparation by OV-M.
Halichoeres nicholsi. Halichoeres nicholsi was described by Jordan and Gilbert (1882), from a single specimen collected at Socorro in 1880 by the USS Hassler and named after the captain of the ship during that cruise. Subsequently 5 other names were synonymized with H. nicholsi by Bussing (1987): Halichoeres sellifer Gilbert, 1890 from Clarion, Halichoeres macgregori from Panama Gilbert and Starks 1904, and 3 from the Galápagos: Halichoeres maculosus Clark, 1936, Halichoeres stictus and Halichoeres stigmasepia, both by Fowler (1944). There are consistent, pronounced coloration differences between all 3 color phases (juvenile, IP adult, TP adult) of the RNP population and those phases from populations on the mainland between Mexico and Ecuador and at the Galápagos (Supplementary material: File S10). These color differences are evident in underwater photos from Clarion, San Benedicto and Socorro (mainly from the 2022 expedition, with 40 others from iNaturalist) as well as many iNaturalist images from the mainland and Galápagos, plus some taken on the mainland in 2023, and at Galápagos in 2024 by AME and CJE: juveniles in the RNP almost invariably (35 of 38 individuals) have red- brown dark markings on a white body, while those from the mainland and Galápagos (and 1 from the Revillagigedos) typically have blackish to chocolate-brown dark markings on a white background (n = 78), and rarely red-brown markings (n = 3). IP adults from the mainland have white and yellow bodies with a blackish mid-lateral stripe along the body and a strong black vertical bar above that stripe under the center of the spinous dorsal fin (n = 187). Those from the RNP invariably (n = 32) have reddish brown bodies and fins, with or without a poorly developed dark vertical bar under the spinous dorsal fin. All terminal phase individuals at the RNP (n = 25) have a pale bluish- white body with an indistinct grey bar at mid-body and a large yellow patch covering the cheek and operculum. Terminal phase fish from the mainland also have bluish- white bodies, plus a broad black bar on the upper 2/3 of the body under the center of the spinous dorsal fin, with a yellow blotch at the front edge of that bar, and blue spots on the cheeks. Only 1 of 124 TPs from the mainland, at Cabo San Lucas, had a color pattern intermediate between those 2 TP color patterns, the remaining 123 exhibited the non-Revillagigedo pattern.
DNA barcoding by ODD (unpublished data) shows that the 27 specimens from Socorro and Clarion form 1 haplogroup and 37 fish from the mainland plus 27 from Galápagos form another, although those 2 haplogroups only differ by a p-distance of 0.23%. Evolutionary change in color patterns seems to proceed more rapidly than change in mtDNA genetic markers (Craig et al., 2006) and, in labrids, color differences are often the most pronounced morphological features separating sister species. The consistent, marked differences in coloration (morphology) between those 2 groups, in combination with that small genetic difference, indicates that the Revillagigedo population is a local endemic, named H. nicholsi.
Lythrypnus cf dalli. This species was discovered and collected at Clarion Is. during the 2022 expedition. Based on its color pattern it is a close relative of Lythrypnus dalli, which has not been recorded in the RNP, although there are some differences in color patterns between the 2 (Supplementary material: File S11). A limited number of DNA barcodes of this species differ by 3.5% from barcodes of L. dalli from Peru and 2.7 % from barcodes from each of the allopatric populations of L. dalli found in California/northern Baja California and in the Gulf of California (ODD and BV, unpublished data), which are known to be well separated genetically (Bernardi et al., 2003). These genetic and morphological differences support the view that the Clarion population likely is an RNP endemic.
Pareques sp. There are 5 species of Pareques currently known from the TEP, 4 named and 1 undescribed (Chao, 1995). They include 3 with very similarly colored plain brown adults: Pareques viola from mainland Nicaragua to Peru, Pareques perissa, endemic to the Galápagos islands, and an undescribed species (Pareques species A; see Robertson & Allen, 2024) from mainland Mexico. The coloration of adults of the Pareques sp. in the RNP is very similar to that of adults of both P. viola and P. species A. from mainland Mexico. In his review of the TEP members of the genus McPhail (1963) only dealt with P. viola among those 3, which he considered to occur in mainland Mexico and the RNP as well as the southern TEP. DNA barcodes of 2 specimens from San Benedicto (from BOLD) and 2 from Clarion (collected by the 2022 expedition) differ by 4.82%, on average, from P. viola and 10.1% from P. species A, with P. viola differing by 9.5% from P. species A (see also the molecular phylogeny based on CO1 in Carvalho-Filho et al., 2022). This result is surprising in that the sister to the RNP population is far more distant from the archipelago (2,800 km) than is the population of P. species A., in Mexico (400 km), although this geographic relationship parallels those seen in B. clarionensis and A. herrei. Based on these levels of genetic difference among those Pareques populations we class the RNP population as probably endemic to those islands.
Tomicodon sp. See species account for T. absitus in endemics section. This species, found only at Clarion, is genetically highly divergent (p-distance separation 19.5%) from T. absitus at Socorro.
Xyrichtys sp. A. The evidence for the presence of this species (Robertson & Allen, 2024) are relatively low- quality photographs of live individuals at Socorro, some from the 1980s and others, by author CC, from 2015. Examination of specimens are essential for confirming the taxonomic identity and status of this species vis a vis RNP endemism. Both 2016 inventories regarded it as an RNP endemic and here we consider that it likely is an undescribed endemic species.
Potential RNP endemics species accounts
Aseraggodes herrei. This small, highly cryptic, sand- living flatfish was described from a single specimen collected in the Galápagos. There evidently is a population in the RNP as it has been collected at Clarion, 3 specimens in 1971, LACM, and Socorro, 2 specimens in 1977, SIO; plus 5 specimens taken at Clarion by the 2022 expedition. ODD and students collected 2 at Clarion and 1 at San Benedicto during their multiyear collecting activity. This species is evidently largely restricted to living at the offshore islands and is also known from the Galápagos, Cocos and Malpelo, a minimum of 2,800 km from Socorro. Single SIO specimens collected in each of 1961 and 1977 at the southeast tip of Baja California might represent waifs from the RNP population. Given the great distance separating those northern and southern TEP populations it is possible that they are sufficiently genetically isolated to represent separate species, with the northern one endemic to the RNP.
Bryx clarionensis. Bryx veleronis was described by Herald (1940) based on 10 specimens collected in the Galápagos (holotype location) in 1938, and 18 individuals obtained at Clarion in 1934, the latter apparently collected by benthic trawl. Subsequently, B. clarionensis was described by Fritsche (1980) from 8 specimens dip-netted at a nightlight at southeast Clarion in 1955. He compared those specimens with 7 of the Clarion specimens of the same size range that Herald used to describe B. veleronis and separated the 2 groups based on the lengths of their snouts relative to head length: distinctly longer in B. clarionensis than in B. veleronis, with no intergradation. He included both species in the RNP fauna. Dawson (1985) synonymized B. clarionensis with B. veleronis, stating: “The recently described B. clarionensis, based on planktonic specimens and distinguished only be a somewhat longer snout, is here considered conspecific with B. veleronis”. While Fourriére et al. (2016) included only B. veleronis in their RNP inventory, Del Moral-Flores et al. (2016) listed both species, with B. clarionensis as an RNP endemic.
We collected 30 specimens of Bryx during the 2022 expedition by dipnet at a nightlight at southeastern Clarion, where Fritzsche’s (1955) specimens were collected. The range of absolute snout lengths in a sample of 9 of those 2022 fish was much wider than reported by Fritsche for both species at Clarion, with intergradation of sizes rather than a dichotomy that he used to separate them. The DNA barcodes of 8 of those fish include 3 closely related haplotypes (0.3% separation; see Bryx), with the individuals with the shortest and longest snouts having the same haplotype. These results support Dawson’s (1985) conclusion that there is only 1 species in the RNP and here we follow that publication, which is the most recent general review of the taxonomy of TEP pipefishes, in calling that population B. veleronis with B. clarionensis as a synonym.
Herald (1940) noted that the 18 Clarion B. veleronis he examined had more dorsal-fin rays (25-28) than the 13 fish from the Galápagos, Costa Rica and Colombia (21-24), while Fritsche (1955) recorded that the 8 B. clarionensis from Clarion also had 25-28 dorsal fin rays. However, dorsal fin rays of the 9 2022 fish ranged from 23 to 26, demonstrating some overlap between the RNP population (by 4 of 35 individuals) and those from the southern TEP. Outside the RNP B. veleronis is known only from mainland Costa Rica to Colombia and the southern oceanic islands (Cocos, Galápagos and Malpelo, but not Clipperton). Hence the RNP population may well be sufficiently isolated from those other populations (by ~ 2,800 km or more) to represent an endemic species, called B. clarionensis. Since genetic data from those southern populations of B. veleronis currently are lacking, the status of the population in the RNP remains to be resolved.
Coryphopterus urospilus. This species, the only member of this neotropical genus in the TEP, is widely distributed throughout the region, from southern Baja California and the Gulf of California to Peru, and all the offshore islands except Clipperton. It is common at the 3 major islands of the RNP. ODD (unpublished data) DNA barcoded 31 individuals from the RNP, as part of a region- wide study. Twenty-five of those belong to a haplogroup restricted to the RNP, with that population separated by an average of 0.6% from other haplogroups found on the mainland and the Galápagos. Species not accepted as RNP endemics We reviewed data available for 6 species considered to be RNP endemics by Del Moral-Flores et al. (2016), one more considered endemic in both 2016 inventories and one that is abundant throughout the RNP islands and much less common on the mainland that was not included among the endemics in either 2016 inventory. The species accounts of those 8 below indicate why we concluded that none are RNP endemics.
Doryrhamphus paulus. Doryrhamphus paulus was described by Frische (1980) from specimens collected in the RNP, who used the name D. melanopleura (type locality eastern Indian Ocean, described in 1858) for the population in the remainder of the TEP and the rest of the Indo-Pacific. Dawson (1981) in turn called the RNP population Doryrhamphus excisus paulus, using Doryrhamphus excisus excisus for the population in the rest of the TEP and much of the Indo-Pacific. Del Moral- Flores et al. (2016) referred to D. paulus as an endemic RNP species, while Fourriére et al. (2016) included it as a non-endemic subspecies of Doryrhamphus excisus. The west and central Pacific populations most recently reverted to the name D. melanopleura, which is also used for non- RNP populations in the TEP by Fricke et al. (2026). DNA sequencing (CytB) of 27 specimens by ODD (unpublished data) from the RNP, plus 50 elsewhere in Mexico, Costa Rica, Ecuador and the Galápagos shows that there are 3 discrete haplogroups in the TEP (Lessios & Robertson, 2006). While 1 haplogroup includes the great majority of RNP specimens, a few individuals from another major haplogroup also occur in the RNP and more than 1/3 of the Galápagos population also belongs to that “RNP” haplogroup. Thus, while there is evidence of partial isolation of the RNP population, the level of haplotype sharing with the Galápagos indicates that there likely is no RNP endemic population. To complicate matters further, Doryrhamphus californiensis (Gill, 1862) was described from Baja after D. melanopleura but before D. paulus and may be either a synonym of D. melanopleura (Frische, 1980) or taxonomically valid (Stiller et al., 2022). Until this complex taxonomic situation in the TEP is formally resolved by further study, we refer to the entire TEP population as D. melanopleura.
Epinephelus clippertonensis. Epinephelus clipperto– nensis was listed as a Revillagigedos endemic by Del Moral-Flores et al. (2016) but not Fourriére et al. (2016). However, while it is common in the RNP it also has a resident population at Clipperton Island, its type locality, and likely one at southern Baja as well. Hence it is not endemic to the RNP.
Gobiesox aethus. See species account in unaccepted- species section.
Halichoeres adustus. This species was described by Gilbert (1890) based on 3 specimens collected at Socorro in 1889. It was listed as an RNP endemic by Del Moral- Flores et al. (2016) but not Fourriére et al. (2016). However, since it is also known from the mainland between the mouth of the Gulf of California and Colombia, as well as Isla del Coco and the Galápagos (Bussing, 1987; Victor, Grove et al., 2024), there is no evidence that the RNP population is endemic. Interestingly, this species appears to be relatively rare in the RNP, with a single individual photographed during the 2022 expedition, with only 2 images in iNaturalist from 2025, and 1 taken by AAB in 2025. There also are very few museum specimens. Hence, we consider its residency status in the RNP to be uncertain.
Kyphosus lutescens. See species account in unaccepted-species section.
Myrichthys pantostigmius. Although Del Moral- Flores et al. (2016), but not Fourriére et al. (2016), included this among the RNP endemics, it is also common enough at Clipperton to be listed as a resident there by Allen and Robertson (1997). Hence, it is not a Revillagigedos endemic. Myrichthys pantostigmius is the only member of the genus with a confirmed presence in the RNP, while reports of Myrichthys xystrurus (as Myrichthys tigrinus) are misidentifications of that species.
Stegastes leucorus. Neither Del Moral-Flores et al. (2016) nor Fourriére et al. (2016) list this species, which is the most common member of its genus throughout the RNP, as an endemic. Other members of its clade are endemics at other TEP offshore islands: Stegastes baldwini at Clipperton and Stegastes beebei at the Galápagos, Cocos and Malpelo. The latter also has scattered isolated mainland records between Costa Rica and Peru. Stegastes leucorus has the largest and most widespread set of mainland records of any potential Revillagigedo endemic, with about 60 primary source records in GBIF, scattered from southern Mexico to southern California. Author CS-O has observed multiple small aggregations of 4-5 adults and also juveniles in the Cabo San Lucas area, which indicates that there likely is a small self-sustaining population on the mainland. Hence, we do not regard it as endemic to the RNP.
Thalassoma virens. This is another species that was also listed as a Revillagigedos endemic by Del Moral- Flores et al. (2016) but not Fourriére et al. (2016). However, Allen and Robertson (1997) recorded it as a resident at Clipperton, and it also occurs around the tip of the Baja Peninsula (Victor et al., 2001). Hence it is not endemic to the RNP.
Deep-reef species
Species of reef-associated fishes in the Greater Caribbean that are mostly or entirely restricted to mesophotic, and greater depths represent about 15% of the regional reef-associated fauna there, and include regional endemics (Robertson et al., 2022). There is information on such species for Cocos and Galápagos, which have had the most intensive taxonomic research involving deep-reef fishes of any sites in the TEP. Table 5 provides a list about 16 such species in the RNP, none of which are endemic to the archipelago.
| Family | Species | Habitat | Resident | Fourriére et al. (2016) | Del Moral-Flores et al. (2016) | Aburto-Oropeza et al. | Giddens et al. (2019) | Hollarsmith et al. (2020) | This study |
|---|---|---|---|---|---|---|---|---|---|
| Chaetodontidae | Prognathodes falcifer | Open reef | Yes | Yes | Yes | Live image | Live image | ||
| Congridae | Paraconger similis | Soft bottom | uncertain | Yes | Yes | Confirmed | |||
| Cyclopsettidae | Citharichthys sp. | Soft bottom | uncertain | New data | New data | Live image | |||
| Epinephelidae | Hyporthodus cifuentesi | Open reef | Yes | Yes | Yes | Yes | Yes | Live image | Live image |
| Holocentridae | Pristigenys serrula | Open reef | Yes | Yes | Yes | Live image | Live image | ||
| Latilidae | Caulolatilus affinis | Reef & sand | uncertain | Yes | Yes | Confirmed | |||
| Latilidae | Caulolatilus princeps | Reef & sand | uncertain | Yes | Yes | Yes | Accepted | ||
| Lutjanidae | Lutjanus peru | Open reef | Yes | Yes | Yes | Yes | Yes | Live image | |
| Ophidiidae | Brotula ordwayi | Reef | Yes | Yes | Yes | Live image | |||
| Pomacentridae | Chromis alta | Open reef | Yes | Yes | Yes | Live image | Live image | ||
| Priacanthidae | Cookeolus japonicus | Open reef | Yes | Yes | Yes | Confirmed | |||
| Priacanthidae | Priacanthus alalaua | Open reef | Yes | Yes | Yes | Confirmed | |||
| Serranidae | Serranus aequidens | Soft bottom | uncertain | Yes | Not listed | Confirmed | |||
| Scorpaenidae | Pontinus vaughani | Open reef | Yes | Yes | Yes | Live image | |||
| Scorpaenidae | Scorpaena afuerae | Reef & sand | Yes | New data | New data | Live image | |||
| Triglidae | Bellator loxias | Soft bottom | No | Yes | Yes | Accepted |
Discussion
Accepted species
The assessments made of 364 potential members of the RNP reef-associated fish fauna indicate that only 234 (64.3%) are validated members, a number that is lower by ~ 12% than the number of the same set of taxa accepted by the “two 2016 inventories”. Among that set of 234 species, 187 are known from Clarion, 126 from Roca Partida, 160 from San Benedicto and 201 from Socorro. These numbers contrast to the number of reef-associated species (178, 155 of which are in our accepted group) listed by Castro- Aguirre and Balart (2002) and numbers of (all types of) species they listed at the 3 main islands: 80 for Clarion, 15 for San Benedicto and 111 for Socorro. All of the reef- associated fishes included in that 2002 checklist are also included in the 2016 inventories and among the species assessed here. The numbers arrived at here also represent substantial increases in the number of species listed by the 2016 inventories: 3.3% more than all (reef-associated and non-reef) species listed for Clarion by Del Moral- Flores et al. (2016) (Fourriére et al. [2016] did not provide individual island data), 98% more at Roca Partida and 26% more at San Benedicto. Although the number in Socorro is 6.5% less than the total number in Del Moral-Flores et al. (2016) that is due to a combination of the abundance there of non-reef species and our rejection of various species they listed as part of the fauna (2016) (see below). That list accepted only 155 of the 201 reef-associated fishes (i.e., 22.9% fewer) we include with validated records from Socorro.
Among the 19 accepted species not mentioned in both 2016 inventories, only 3 (P. multifasciatus, Diplobatis ommata, and the probable endemic Chriolepis sp.) had been collected prior to the publication of those inventories, while all the others are more recent additions. Those 19 include 7 species added by the 2022 expedition that were not in either 2016 inventory and represent new additions to the fauna: C. alepidota, Gymnothorax verillii, Lutjanus aratus, O. clippertonensis, and T. purpureum, plus 1 new endemic (H. sanchezi) and 1 probable endemic (L. cf. dalli. at Clarion). Three of those represent new residents: O.clippertonensis and the 2 endemics.
Analyses of the functional structure of reef-fish faunas based on simple faunal inventories typically include all members of the fauna (e.g., Bender et al., 2017; Dubuc et al., 2023; Ferrari et al., 2023; Palacios-Salgado et al., 2019). We suggest that at this “occurrence” level the most appropriate set of species for functional-structure analyses is the residents, although local inventories often do not provide information on residency. Residents represent the ecologically important sector of the fauna and non-residents are too rare as individuals to have an ecological impact on the residents or other elements of the local ecosystem. The residents among the accepted set for the RNP differ from those in the 2 2016 checklists. Three species in 3 genera have been removed (G. aethus, K. lutescens, and Prionurus punctatus) and 10 species in 10 genera added: O. clippertonensis, P. multifasciatus, D. ommata, Quassiremus evionthas, S. afuerae, and C. janthinoptera, plus 4 new actual or probable endemics Tomicodon sp., Chriolepis sp., L. cf. dalli, and H. sanchezi. However, the effects of that change in this small percentage (< 5%) of residents on the functional structure of the fish assemblage would be much smaller than the effects of the removal of ~ 20% non-residents from the assemblage. The results of functional analyses would also have differed without the removal of the unaccepted third of the species from a wide diversity of genera and families that we assessed (see below) or the rejection of 20.4-22.1% of the species accepted by the 2 2016 checklists (see below). Photographic database Species with photographs represent 59-75% of those known from each island and images from various sources presented here document 77.4% of the 234 accepted species. This represents a solid start to more complete documentation of the fauna using taxonomically diagnostic images, which unequivocally demonstrate occurrence, and, in some cases, provide information about population status. Those numbers also demonstrate the value that citizen scientist contributions can make to research on faunal inventories, some as coauthors of this paper and others as depositors of taxonomically useful images in iNaturalist.
Unaccepted species
Our assessments of the 364 reef-associated fishes that potentially are members of the RNP fauna indicate that there are erroneous records of 121 (33.2%) of the assessed species, 19% of the assessed genera and 18% of the assessed families (Table 2). Those errors are due to a variety of reasons, with most arising from misidentification or inability to identify (38.0%), a lack of vouchers (30.6%) or incorrect location data (37.2%). Loss of museum specimens in 18.0% of those species means that their occurrence is unresolvable using existing information. Multiple types of problems were present in 33.1% of those species. Of those 121 species, 27 were included in the 2002 checklist, 32 were not mentioned in either of the “two 2016 inventories”, 52 had been accepted by both and another 13 accepted in one 2016 inventory but not by the other. Eight species were accepted on one 2016 inventory but rejected on the other and only 3 species had been rejected on both those inventories. Re- examination of museum specimens of 72 species showed that 38% of those species had been misidentified or, in a few cases, were unidentifiable, in some species involving multiple individuals from multiple museums. Those misidentified and unidentifiable cases involved older records of specimens collected in the 19th to mid-20th century, highlighting the necessity to physically review specimens in older collections. Reviewing metadata of old collections is also important, as experience with the 1889 Albatross collections has shown. The level of rejection of potential members of the RNP fauna is distinctly higher than in other studies of fish faunas that culled erroneous and unsupported records (Borg et al., 2023; Mundy, 2005; Mundy et al., 2010; Victor, Grove et. al., 2024). This reflects the relative importance of the previously unassessed inventory of old records that have been found to be erroneous in the present case. However, among those 121 excluded species are a number whose proximity to the RNP on mainland Mexico could allow them to recruit to the RNP and be included as validated members of the fauna in the future.
Endemic species and the endemism rate
The “two 2016 inventories” and the present study differ in terms of how many species they regard as endemics or probable endemics and which species those are, due to a combination of factors. Those include new information about the genetics and morphology of species long known from the RNP (A. perezponcedeleoni, H. nicholsi, B. ramosus, Pareques sp.), the recent discovery of new additions to the RNP fauna (H. sanchezi, L. cf dalli, Tomicodon sp.), synonymization (G. aethus and K. lutescens), rediscovery of an unnoticed, probably endemic species (Chriolepis species) and the presence (E. clippertonensis, H. adustus, M. pantostigmius, and T. virens) or absence (L. insularis, H. clarionensis, S. redemptus) of resident populations of species outside as well as inside the RNP. The 3 main islands have similarly high numbers of endemics and probable endemics. Most actual and probable endemics are present at more than 1 of the main islands and are present at Clarion as well as the eastern islands (Table 4). All but 2 species are present at multiple islands, with 2 likely restricted to Clarion and another 2 restricted to multiple eastern islands, a reflection of the fact that Clarion is as isolated from the main eastern islands as they are from Baja California.
Fourriére et al. (2016) estimated the reef-fish endemism rate to be 5.5% of 235 species or 4.8% of the expanded reef-fish inventory of 271 from that paper that includes various other taxa and ecotypes that we included as reef- associated, as did Fourriére et al. (2017) for Cocos. Del Moral-Flores et al. (2016) in turn estimated 7.1% of the 366 species of all types that they regarded as validated, equivalent to 9.8% of the 268 species of those that are reef- associated. However, they overestimated the number of endemics (see above) and taking that into account reduces their endemism rates to 4.9% and 6.7% respectively. Other estimates of RNP endemism rates include 9-10% of a fauna of ~ 100 species by Briggs (1974) and 8% of 212 species by Robertson and Cramer (2009), and Briggs and Bowen (2013). The 16 named endemics plus 7 probable endemics identified here produce an estimate of 9.8% of the entire 234- member reef-associated assemblage accepted here. However, endemics have evolved in response to local, long-term conditions in the RNP, including the persistent presence in abundance of other residents. Since at any time non-residents are present in very small numbers and, that presence often is transient, that group of species is ecologically irrelevant to the community-level evolutionary processes involved in the development of endemics. Hence, we suggest that the most relevant part of the assemblage for determining the endemism rate is the resident species. Our data indicate that the endemism rate of known endemics is 8.5-9.6% of the residents, and 12.2- 13.9% of the residents if probable endemics are included.
Two other TEP offshore islands have comparably comprehensive data about endemism rates: Cocos Island and the Galápagos. For Cocos, 13 species of endemics represent 3.7% of the entire reef-associated fauna of 355 species (Fourriére et al., 2017; who do not present information on the residency of species in their inventory). For the Galápagos, 43 endemics represent 10.2% of the entire reef-associated fauna of 432 species and 12.2% of the members of that fauna that are residents (Victor, Grove et al., 2024). However, while the rate of endemism in the RNP is similar to that of the current estimate for the Galápagos, we think the RNP rate will rise further as more research is done on the integrative taxonomy of members of the RNP fauna and may also do so for Cocos and Galápagos with more such research there as well.
Deep-reef fishes in the RNP
Such fishes are members of typical reef-fish families that have depth ranges restricted largely or entirely to mesophotic and greater depths that are below normal scuba limits (Baldwin et al., 2018; Pinheiro et al., 2019). In the Greater Caribbean intensive sampling with crewed submersibles has shown that deep-reef fishes constitute substantial proportions of both regional and (well-studied) local faunas (Robertson et al., 2022). A total of 16 such species are currently known from the RNP. They include 14 species listed by the “two 2016 inventories” plus 2 recently found through examination of images collected by a submersible (by ABB) and BRUVS (by CAS-O) in 2016. The 15 of those that are named represent 6.3% of the accepted species, 10 are residents and none are endemics. In contrast, at Cocos Island, 46 (13.0%) of the 355 reef- associated fishes (Fourriére et al., 2017) and the Galápagos, 59 (13.7%) of 432 reef-associated fishes are deep-living species (Victor, Grove et al., 2024). In both cases they include endemics (3 at Cocos and 6 at the Galápagos) as well as other residents. Those 2 sites have a much longer history of more intensive taxonomic research on their deep faunas than does the RNP. To date there has been no research specifically aimed at thoroughly documenting the deep-reef fish fauna of the RNP, particularly any small, cryptic species like those that feature in the deep- reef faunas of Cocos and the Galápagos. Hence, it is quite probable that more research will yield not only more deep- reef additions to the RNP fauna, but also more (deep) endemics.
How many valid reef-associated fishes are known from the RNP?
Fourriére et al. (2016) accepted as validated 235 species of reef fishes, plus another 36 species they did not class as reef fishes but which we include in that group. Fourriére et al. (2017) also included as reef fishes in their inventory of Cocos Island reef fishes only part of the set of species we class here as reef-associated. Fourriére et al. (2016) discounted only 13 species as doubtful or unconfirmed. Our assessment agrees with their 13 discounts but does not accept 5 of the species they did accept. Thus, among the 271 species Fourriére et al. (2016) included in their checklist (of both reef and non-reef fishes) we accept as validated only 213, 78.6% of those they accepted and 91.0% of those we accept as validated. Del Moral- Flores et al. (2016) in turn accepted only 210 (89.7%) of the 234 species of reef fishes we included as validated, accepted another 58 species we rejected, and discounted a further 17 species we also did not accept. Thus, the net number of species among the 268 included that are also included in our set of accepted species is 78.9% of those they accepted. Neither of those “two 2016 inventories” mentioned another 37 species that we assessed but did not accept or 19 other species that we did accept and only one of those 2 inventories accepted another 6 species that we also accepted.
Our assessment indicates that there are validated records for 234 reef-associated fishes and that we do not accept records of another 121 species for various reasons. This set of validated species is 16-17% smaller than the sets of species included in the “two 2016 inventories”, but ~ 11% larger than numbers of such species in those inventories that we regard as validated.
Most previous sampling in the RNP has been directed at rocky reefs in shallow (scuba depth or intertidal) water. However, that sampling failed to find a substantial number of species that use that habitat and were collected and photographed subsequently to the publication of the 2016 inventories. There are 16 species of deep-living reef fishes currently known from the RNP (Table 5). These, which include both hard-reef species and those living in sand habitat, constitute 6.8% of the reef-associated fishes in the RNP fauna as currently known. The relative abundance of such deep-reef fishes is distinctly higher at other TEP offshore islands with much better sampled deep habitats, 13% of those at Cocos (Fourriére et al., 2017) and 14.7% at Galápagos (Victor, Grove et al., 2024). In addition, while the former both have deep-living endemics, none are currently known from the RNP.
Insufficient collecting attention has been directed at RNP habitats other than rocky reefs, including sand and rubble bottoms in both shallow (scuba-depth) areas as well as deep-reef hard-bottom and sand-rubble habitats. Rhodolith beds are a pantropical habitat often found associated with reefs and are known to support diverse faunas of reef-associated fishes (Anderson et al., 2023). In the TEP t rhodolith beds occur from Mexico to Panama on the mainland, as well as Cocos, the Galápagos and the RNP (Diaz-Licona et al., 2025). Using a small ROV, Hollarsmith et al. (2020) found extensive rhodolith beds at Clarion at 40-80 m, in which they recorded 25 species of reef fishes, with habitat variation produced by differing fleshy algal assemblages associated with those beds at different parts of the island. A recent study of shallow (2- 25m depth) rhodolith beds at Cocos Island (Diaz-Licona et al., 2025) listed 37 species of reef fishes living in them. Sampling effort, particularly towards small cryptobenthic species, should be directed at that habitat at all the RNP islands. To date, most research has been focused on the 3 eastern islands of the RNP, due to their greater accessibility. Additional field and laboratory research on integrative taxonomy needs to be made to compare and contrast the faunas of the eastern islands (Roca Partida, San Benedicto and Socorro) with that of Clarion. Given the existence of endemics restricted to either Clarion or the eastern islands, further work on documenting the Clarion fauna likely would yield additional species and local endemics. Due to these sampling biases, we think that there are more, quite possibly substantially more, unknown reef-associated fishes in the RNP that more extensive sampling of such habitats would reveal, including the discovery of deep-living endemics.
Concluding remarks
There are 2 aspects to the construction of this new inventory of RNP reef-associated fishes: reassessment of existing information and the addition of new information. The assessment of the potential inventory of RNP reef- associated fishes described here produced marked changes in the validated inventory, with non-acceptance of 1/3 of the potential members of the fauna that were indicated by old information. This demonstrates that detailed examination of the variety of primary sources of data is essential for producing reliable inventories and culling erroneous and unsupported records. Grey literature reports and ichthyological publications that do not provide details of, or assess the validity of, voucher information and simply recycle previously published information only tend to produce inflated, unreliable inventories. Our analysis has shown how common unsupported records can become in inventories. Older museum records, particularly those from research programs that collected at multiple locations scattered over a region on multiple occasions, can have significant error rates due to a lack of rigorous record keeping, loss of original associated metadata and fragmentation and transfers of parts of old collections among different collections at different times. Misidentifications of specimens in old collections reflect the lack of the range of reliable information that is now available: comprehensive taxonomic studies that are identifiable through Fricke et al. (2026); comprehensive online databases from museums that provide details about their specimens and their provenance; and museum curators who have the time to check identification of specimens and their provenance metadata; and diving imagery produced by citizen scientists, such as those from RNP activity by iNaturalist contributors and by members of the 2022 expedition. Participation of citizen-scientist photographers in expeditions and as coauthors of arising publications to which they make significant contributions will provide recognition of the value of their efforts (Mason et al., 2025).
Expanding the base of such information through recent field and laboratory research and analyses has also made significant contributions to understanding what species are in the RNP and what their status is with respect to residency and insular endemism. Citizen scientists who contribute images, for example to iNaturalist, have made and will continue to make invaluable contributions to that effort. New genetic evaluations of individuals and populations of reef-associated fishes in the RNP that we reported and the first results of eDNA studies have demonstrated their utility for additions to the inventory through species identification and assessment of the level and geographic pattern of endemism within the RNP as well as between the RNP and the mainland and elsewhere in the TEP. The isolation of Clarion from the eastern islands of the RNP, which has resulted in it being under-sampled during this century, has produced some microendemics, as demonstrated by the few genetic data mentioned above. The number of members of the fauna, including endemics, newly recorded by a single, short expedition in 2022 clearly demonstrates that there is much still to learn.
Fourriere et al. (2016) statistically estimated the size of the RNP reef-fish fauna, and concluded that it was well documented, with only ~ 9 species likely present in addition to the 235 they listed. However, assessments of those 235 species led to us including only 177 of them in the group of accepted species presented here, to which we added 21 species not mentioned in their study, 7 of them newly discovered during the 2022 expedition, plus 36 species that Fourriere et al. (2016) listed as non-reef types. Further research likely will show that some of the 121 species that we did not accept due to data inadequacies and some of the 9 species requiring confirmation are valid members of the fauna. In addition, there are reasons to believe that other species of reef-associated fishes await discovery in the RNP. There are sampling deficits of particular habitats (e.g., rhodolith beds and other deep- reef habitats, which could be sampled by technical divers); comprehensive review of various ROV, BRUV and submersible videos of shallow and deep reef fishes is lacking; there have been highly uneven sampling efforts directed at different ecotypes of fishes at different times during the past 125 years (e.g., no use of ichthyocides to collect hidden cryptobenthic species in subtidal areas for many decades); and there has been much less comprehensive sampling at Clarion than at the eastern islands. Many species currently accepted as part of the RNP fauna evidently are vagrants. Ocean currents that impinge on the archipelago carry fish recruits there from Baja California, the Mexican mainland and southern parts of the TEP, as well as the central Pacific. The influx of currently undocumented vagrants can be expected to continue indefinitely, gradually increasing the number of species confirmed at the RNP. While the present size of the RNP fauna undoubtedly is larger than the 234 species documented here, the question of how much larger would best be addressed through research aimed at an array of sampling gaps and at genetic relations among populations of broad range of species found at both Clarion and the eastern islands of the RNP.
Acknowledgements
Thanks to Dora María Sierra Palma (owner of the Liveaboard, Quino El Guardián, which supported the 2022 expedition), Susan Long of Mexicanliveaboards. com and the crew of the Quino El Guardián during that expedition. H. J. Walker, P. A. Hastings, and J. M. Kim provided help with specimen identification and analyses of museum records at SIO; D. Catania assisted regarding the existence of specimens at CAS, N. Bailly identified some UBC specimens and facilitated loans of others to the LACM for review. C. Lowe helped with the identification of shark specimens at Long Beach Museum, D. Long with identifications of elasmobranchs and A. Driskell (USNM) for help with DNA data. Arturo Ayala- Bocos would like to thank the Undersea Hunter Deep Sea submersible and its pilot, Diana Benito, for sharing some videos used here. Angélica Tamayo provided the image of P. teira she took at Socorro in 2025, and Jordan Hollarsmith shared images from her RNP ROV study. Various people helped obtain the permission to collect during the 2022 expedition: Alberto Tirado Arámburo, Universidad Autónoma de Baja California Sur (UABCS); Gustavo Cruz Chávez, UABCS; Everardo Mariano Meléndez (Conanp); Josué Melesio Tiscareño Villorin (Conanp); Luz Eréndida Frías Hernández (Conanp) and Georgina Guadalupe Villavicencio Rousseau (Conanp). The Mexican Navy provided generous logistical support for collecting activities of the research group of ODD at Socorro and Clarion islands between 2015-2023, ODD extends a special thanks to all the people from Universidad Michoacana de San Nicolás de Hidalgo who helped with fieldwork between 2015-2023. Samples were collected under the permits SGPA/DGVS/02920/15 of 18 March 2015 and its extension SGPA/DGVS/12333/15 of 28 October 2015 granted to JEMG and PPF/DGOPA-035/15, PPF/DGOPA-065/21, PPF/DGOPA-085/22, and SGPA/ DGVS/005723/18 to ODD. Collections and photography during the 2022 expedition were done under permit PPF/ DGOPA-099/22 to Carlos Armando Sánchez Ortiz. We thank a reviewer for useful comments on the ms.
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