
Taxonomía y Sistemática

Ecología






Genetic diversity and phenotypic variation in a parasitoid wasp involved in the yucca – yucca moth interaction
C. Rocío Álamo-Herrera a, María Clara Arteaga a, *, Rafael Bello-Bedoy b
a Centro de Investigación Científica y de Educación Superior de Ensenada, Departamento de Biología de la Conservación, Carretera Tijuana-Ensenada # 3918, Zona Playitas, 22860 Ensenada, Baja California, Mexico
b Universidad Autónoma de Baja California, Facultad de Ciencias, Carretera Transpeninsular # 3917, Colonia Playitas, 22860 Ensenada, Baja California, Mexico
*Corresponding author: arteaga@cicese.mx (M.C. Arteaga)
Received: 28 February 2024; accepted: 01 July 2024
Abstract
Tri-trophic interactions between plants, herbivores, and parasitoids are a valuable model for studying how they influence the distribution of genetic diversity and phenotypic variability of the species involved. This study examines the taxonomic, morphological, and genetic diversity of parasitoid wasps involved in the Yucca–Tegeticula interaction on the Baja California Peninsula. We surveyed 35 locations across the peninsula and collected 119 parasitoid wasps. Of these, 114 were adults, while the remaining 5 were in the pupal stage. Our study identified 2 genera of wasps: Bassus sp. (Ichneumonidae; n = 8) and Digonogastra sp. (Brachonidae; n = 111). Moreover, we found moderate levels of genetic diversity within the Digonogastra population across the peninsula. Additionally, this population constitutes a single panmictic group with indications of historical demographic expansion. Phenotypically, we identified sexual dimorphism and variation associated with its different hosts and environmental heterogeneity Digonogastra’s geographical range.
Keywords: Baja California Peninsula; Genetic structure; Host-association; Morphometrics; Tri-trophic interactions
© 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/).
Diversidad genética y variación fenotípica en una avispa parasitoide involucrada en la interacción entre yucas y sus polillas polinizadoras
Resumen
Las interacciones tritróficas entre plantas, herbívoros y parasitoides son un modelo valioso para estudiar cómo influyen en la distribución de la diversidad genética y la variabilidad fenotípica de las especies involucradas. Este estudio examinó la diversidad taxonómica, morfológica y genética de avispas parasitoides en la interacción Yucca-Tegeticula en la Península de Baja California. El estudio se realizó en 35 localidades recolectando 119 avispas parasitoides; 114 adultos y 5 pupas. Se identificaron 2 géneros de avispas: Bassus sp. (Ichneumonidae; n = 8) y Digonogastra sp. (Brachonidae; n = 111). Se encontraron niveles moderados de diversidad genética dentro de la población de Digonogastra en toda la península, constituyendo un único grupo panmítico con indicios de expansión demográfica histórica. Fenotípicamente, identificamos dimorfismo sexual y variación asociada con sus diferentes hospederos y la heterogeneidad ambiental a lo largo de la distribución geográfica de Digonogastra.
Palabras clave: Península de Baja California; Estructura genética; Asociación al hospedero; Morfometría; Interacción tri-trófica
Introduction
Tritrophic interactions between plants, herbivores, and parasitoids have become pivotal to understanding species diversity (Abdala-Roberts et al., 2019; Godfray, 1994; Singer & Stireman, 2005). Parasitoids maintain an antagonistic relationship with insect herbivores by depositing their eggs inside or on them, ultimately leading to the death of their host (Godfray, 1994; Quicke, 2015; Resh & Cardé, 2009). These parasitoids serve as an indirect defense for plants, controlling herbivore population levels (Abdala-Roberts et al., 2019; Cuautle & Rico-Gray, 2003; Heil, 2008). Plants attract parasitoids by emitting chemical signals that indicate the presence of herbivores, which enables parasitoids to locate their prey, thus establishing mutually beneficial interactions (Heil, 2008; Kappers et al., 2011; Takabayashi & Dicke, 1996).
Multiple studies have explored how interactions among organisms affect the genetic and phenotypic variation within species (e.g., Agrawal, 2001; Carmona et al., 2015). For instance, biotic interactions may differ geographically, resulting in local selection processes and differentiation of parasitoid populations (Althoff & Thompson, 2001; Kankare et al., 2005; Stireman et al., 2005). Environmental factors or geographical distances can also determine the distribution of these variations (Althoff, 2008; Lozier et al., 2009; Stireman et al., 2005). For example, the genetic population structure of the wasp Cotesia congregata Say, 1836 (Hymenoptera: Braconidae) is related to the different plant-hosts with which it interacts (Karns, 2009). Conversely, genetic diversity in the parasitoid wasp Eusandalum sp. Ratzeburg, 1852 (Hymenoptera: Braconidae) is primarily associated with its broad geographical distribution rather than the species it interacts with (Althoff, 2008).
The interaction between Yucca Linnaeus plants, moths of the family Prodoxidae and associated parasitoids have been recorded (Althoff, 2008; Pellmyr, 2003). In this tritrophic interaction, the female moth visits Yucca flowers and lays her eggs in the ovary. Subsequently, she pollinates the stigma by depositing pollen, ensuring the formation of fruits that the moth larvae will feed on (Engelmann, 1872). During fruit production, female parasitoid wasps use their ovipositors to lay eggs on moth larvae inside fruits, paralyzing the larvae (Force & Thompson, 1984). The wasp larva feeds on the host, completes its development, and emerges from the fruit as an adult (Althoff, 2008; Crabb & Pellmyr, 2006). The interaction between Yuccas and moths have driven differentiation and diversification processes in the involved species (Althoff et al., 2012; Althoff & Segraves, 2022; Pellmyr & Leebens-Mack, 1999). However, little is known about the third trophic level of this relationship, which consists of parasitoid wasps that interact with the moth (their food source) and the plant (their shelter until hatching).
In the Baja California Peninsula, 3 Yucca species and 2 Tegeticula Zeller, 1873 species are distributed allopatrically. Yucca schidigera Roezl ex Ortgies (Asparagales: Asparagaceae) occurs in the northern part of the peninsula and is pollinated by Tegeticula mojavella Pellmyr, 1999 (Lepidoptera: Prodoxidae). Yucca valida Brandegee (Asparagales: Asparagaceae) is distributed in the central region, whereas Yucca capensis L.W. Lenz, 1998 (Asparagales: Asparagaceae) occurs in the southern part of the peninsula. Both Y. valida and Y. capensis are pollinated by Tegeticula baja Pellmyr, Balcázar-Lara, Segraves, Althoff & Littlefield, 2008 (Lepidoptera: Prodoxidae; Lenz, 1998; Turner et al., 1995). Furthermore, a region of hybrid populations of Y. valida and Y. capensis has been identified, both pollinated by T. baja (Arteaga et al., 2020). However, there are no previous records of parasitoid wasps associated with T. baja or T. mojavella populations in the Baja California Peninsula. This study aims to identify the genera of parasitic wasps associated with Tegeticula species in the peninsula. We also investigate whether the use of different hosts, such as T. mojavella and T. baja, and the environmental distribution of the wasps lead to phenotypic and genetic differentiation in the wasp populations.
Materials and methods
We visited 35 locations in the Baja California Peninsula, following the distribution of the Yucca species and their pollinators T. mojavella and T. baja (Fig. 1A, Table S1). Specifically, we surveyed 12 locations in the northern section of the peninsula, where Yucca schidigera occurs, within forest habitats of Sierra Juárez (N = 6), Sierra San Pedro Mártir (N = 4), and Chaparral (N = 2). We visited 11 locations within the central desert of the peninsula where Yucca valida are distributed. Finally, we collected samples from 12 locations in the south section of the peninsula. Eight locations were in the coastal plains of the Magdalena Plains region, where populations of Y. valida x Y. capensis are found. The remaining 4 locations were in the deciduous lowland forest of the Cape Region, where Y. capensis populations are present.
Each location was visited once between 2013 to 2015, during the fruiting season of the Yucca species. We selected approximately 10 trees per location, gathering 3 to 5 fruits from each tree. The mature fruits were collected directly from Yucca trees and placed individually within 500 ml plastic cups with mesh netting lids. Fruits were transported to the laboratory and stored in rooms at environmental conditions (approximately 25°C and 60% relative humidity). For 2 weeks, we checked each plastic cup daily for adult wasps. All adult wasps that emerged from the fruits were collected and placed in 20 ml glass vials. Following another 2 weeks, we dissected the fruits to obtain wasps pupae. All wasps were preserved in glass vials with 96% ethanol and labeled with locality and host plant species. Adult individuals were observed with a Nikon SMZ745-T stereomicroscope equipped with Lumenera’s INFINITY digital camera, and identified to genus taxonomic level using the dichotomous key provided by Sánchez et al. (1998). Two genera of parasitoid wasps from the Braconidae family were identified (Fig. 2): Bassus Fabricius, 1804 and Digonogastra Viereck, 1912.
Figure 1. A, Localities sampled of Digonogastra sp. in the Baja California Peninsula. The geographical distribution is marked using colors and the host moth species is indicated whit shapes; B, haplotype network. The geographical distribution is marked using colors and the host moth species with lines; C, graph depicting the observed and simulated distribution of paired sequence differences.
Figure 2. The genera of parasitoid wasps sampled from yucca fruits in the Baja California Peninsula. On the upper side is the genus Digonogastra; on the bottom is Bassus.
DNA extraction and molecular marker amplification. We extracted DNA from 119 wasps using the commercial Qiagen DNeasy Blood & Tissue Kit. The sample consisted of 114 adult individuals and 5 in the pupal stage. A fragment of the Cytochrome Oxidase subunit I (COI) marker was amplified via PCR using the universal primers for invertebrates described by Folmer et al. (1994); LCO1490 (5’-GGTCAACAAATCATAAAGATATTGG-3’) and HCO2198 (5’-TAAACTTCAGGGTG ACCAAAAAATCA-3’). The PCR mixture included 5 µl of Buffer (1x), 1.5 µl of MgCL (2.5mM), 0.3 µl of dNTPs (0.16mM), 0.6 µl of each primer at 10 µM, 0.2 µl of Taq polymerase (1 unit), 1 µl of DNA, and 5.8 µl of molecular-grade water, resulting in a 15 µl reaction volume.
The thermal cycler was set up with the following conditions: an initial denaturation at 94 ºC for 5 min, followed by 35 cycles of denaturation at 94 ºC for 1 min, annealing at 50 ºC for 1 min, and extension at 68 ºC for 1 min. A final elongation step was performed at 72 ºC for 5 min. Amplification quality was confirmed using 1% agarose gel electrophoresis. The PCR products were sequenced by SeqXcel (www.seqxcel.com) for further analysis.
Genetic diversity and population genetic structure. The sequences were visualized, aligned, and edited using the BioEdit software (Hall, 1999). Sequences from individuals identified morphologically were submitted to BLAST to confirm the parasitoid genus (Blast.ncbi.nlm.nih.gov). The sample size of Digonogastra sp. (N = 111) allowed for diversity and population structure analysis. The genetic diversity of Digonogastra species was assessed using DNAsp software (Rozas et al., 2003). This involved calculating the number of haplotypes, haplotype diversity, and nucleotide diversity (Nei & Li, 1979; Nei, 1987). To investigate the genealogical relationships among the identified haplotypes, we constructed a haplotype network using the Median-Joining method in NETWORK 5.0 software (Bandelt et al., 1999). We constructed a phylogenetic tree using the haplotypes obtained for Digonogastra sp. to determine whether the detected diversity in the Baja California Peninsula is unique to this region or present elsewhere. We included sequences available in the NCBI from Canada. The genera Alabagrus (Sharkey & Chapman, Unpublished; GenBank: MF361682.1) and Cotesia (Hebert et al., 2016) were employed as outgroups. The tree was constructed using the Maximum Likelihood method, with 1000 bootstrap replicates and the HKY+I substitution model, which showed the best fit to the data (highest AIC value), as determined by the jModelTest program (Darriba et al., 2012).
To assess genetic differentiation of Digonogastra spp. across its geographical distribution in the peninsula, we conducted 3 Molecular Variance Analyses (AMOVA). First, we examined how geographical distances affected the distribution of genetic diversity. We divided the data into 3 categories based on their location in the peninsula: north, center, and south (Fig. 1A). Then, we assessed whether genetic differentiation was due to environmental factors, grouping the data based on their ecoregion of origin. We based our categorization on the ecoregions proposed by Gonzales-Abraham et al. (2010). Lastly, our third analysis examined whether genetic differentiation was related to the host, grouping the data based on the host moths, T. mojavella and T. baja. The ARLEQUIN software (Excoffier et al., 2005) was employed for these analyses.
We evaluated historical demographic changes in the wasp population using the pairwise sequence differences distribution analysis (mismatch analysis) performed in the ARLEQUIN software. The shape of the mismatch distribution is used to infer whether a population expansion has occurred (Rogers & Harpending, 1992). A unimodal distribution indicates population expansion, while a multimodal distribution suggests a stable population size. Additionally, the sum of squared deviations (SSD) is employed to validate the expansion model (Navascués et al., 2006). A significant SSD (p < 0.05) rejects the population expansion model.
Phenotypic variation. Phenotypic variation of the parasitoid wasps was evaluated by measuring 106 adult
individuals of the Digonogastra genus (excluding individuals in the pupal stage). No morphometric analyses were conducted for Bassus specimens due to their small sample size (N = 8). Measurements of the 106 adults were conducted with Infinity Analyze software (Lumenera, Canadá), calibrated in millimeters and verified with a conventional ruler. We measured 18 external morphological characters of the adult individuals (Table 1). Measurements were taken on the left side of the individuals and included the length from the head to the end of the last metasoma segment, the scape width, antenna length, mesosoma (thorax) and metasoma (abdomen) width and length, femur width, total leg length, anterior wing length, and the length of 5 wing vein components: C+SC+R, 1RS, (RS+M)a, 2RS, and r-rs. For females, we also measured the length and width of the ovipositor and the length of the ovipositor apex (Table 1). We calculated the mean, standard deviation, and coefficient of variation of each measured character. A correlation matrix was created using the Pearson Correlation Coefficient (r) between pairs of characters, determining the significance values (p) for each correlation. All morphometric measurements were analyzed with JMP 5.01 software (SAS Cary, New Jersey, USA).
We assessed 3 potential sources of variation for the phenotypic differentiation among parasitoid wasps in the Baja California Peninsula: sexual dimorphism, the ecoregions in which they are distributed, and the host species of Tegeticula moths. We conducted a Multivariate Analysis of Variance (MANOVA) for each factor using all measured characters. We also performed an Analysis of Variance (ANOVA) to determine which trait contributes significantly to phenotypic differences. Each wasp was assigned to an ecoregion based on its geographical origin (Gonzales-Abraham et al., 2010). The wasps were found in 6 ecoregions: Sierra Juárez, Sierra San Pedro Mártir, Chaparral, Central Desert, Magdalena Plains, and Cape Lowland Forest.
Results
We collected a total of 119 parasitoid wasps from 35 locations across the Baja California Peninsula (Fig. 1A, Table S1). Two genera of parasitoid wasps belonging to the Braconidae family were collected: Bassus and Digonogastra. Eight female individuals of Bassus sp. (6.7%) emerged from fruits in 2 different locations. Seven Bassus specimens were collected in Y. valida fruits, while 1 emerged from Y. capensis. All these fruits contained T. baja larvae. In contrast, we collected 111 Digonogastra individuals (93.3%), with 68 males and 38 females from 33 different sites across 32.58° to 23.38° N latitude. Out of these, 57 individuals of Digonogastra sp. emerged from Y. schidigera fruits where T. mojavella was also found. The remaining 54 individuals were found in Y. valida, Y. valida x Y. capensis, and Y. capensis fruits, where T. baja larvae were also present.
Table 1
Mean (M), standard deviation (SD), and coefficient of variation in percentage (CV) of the evaluated morphological traits in female and male individuals of Digonogastra sp. in the Baja California Peninsula. The results of the analysis of variance (ANOVA) performed for sexual dimorphism (S), ecoregions (E), and host (H) are presented with significance levels denoted as follows: p > 0.05 (ns), p < 0.05 (*), and p < 0.0001 (***).
Trait | Female | Male | ANOVA | ||||||
M | SD | CV | M | SD | CV | S | E | H | |
Body length: | 9.03 | 1.29 | 14.24 | 7.34 | 1.47 | 20.04 | *** | * | ns |
Antenna: | |||||||||
Total length | 6.88 | 0.73 | 10.58 | 6.16 | 1.08 | 17.52 | *** | * | ns |
Escapo width | 0.24 | 0.03 | 12.61 | 0.19 | 0.04 | 22.92 | *** | ns | ns |
Leg: | |||||||||
Total length | 6.97 | 0.90 | 12.96 | 5.21 | 1.08 | 20.67 | *** | * | ns |
Femur width | 0.52 | 0.06 | 12.23 | 0.36 | 0.08 | 22.93 | *** | * | ns |
Mesosoma: | |||||||||
Lateral width | 2.17 | 0.32 | 14.54 | 1.61 | 0.34 | 21.11 | *** | ns | ns |
Lateral length | 3.17 | 0.45 | 14.08 | 2.46 | 0.58 | 23.59 | *** | * | ns |
Metasoma: | |||||||||
Lateral width | 2.14 | 0.56 | 26.32 | 1.33 | 0.44 | 33.15 | *** | * | ns |
Lateral length | 4.84 | 0.81 | 16.83 | 4.03 | 0.86 | 21.20 | *** | * | ns |
Wing: | |||||||||
Total length | 8.47 | 0.95 | 11.19 | 6.30 | 1.26 | 19.96 | *** | * | ns |
C+SC+R length | 3.99 | 0.48 | 12.14 | 3.03 | 0.62 | 20.62 | *** | * | ns |
1RS length | 0.26 | 0.04 | 15.15 | 0.21 | 0.05 | 23.26 | *** | ns | ns |
(RS+M)a length | 1.00 | 0.13 | 12.46 | 0.70 | 0.15 | 21.33 | *** | * | ns |
2RS length | 0.69 | 0.10 | 13.98 | 0.51 | 0.10 | 18.78 | *** | * | ns |
r-rs length | 0.29 | 0.05 | 17.53 | 0.21 | 0.05 | 21.47 | *** | * | ns |
Ovipositor: | |||||||||
Total length | 8.37 | 1.07 | 12.84 | NA | NA | NA | NA | ns | ns |
Lateral width | 0.06 | 0.01 | 10.54 | NA | NA | NA | NA | ns | ns |
Apex length | 0.36 | 0.05 | 14.22 | NA | NA | NA | NA | ns | ns |
Genetic diversity and population genetic structure. We obtained a 632 bp sequence for each of the 8 individuals of Bassus sp. These sequences did not exhibit site variability, defining them as a single haplotype. Further analyses were not conducted due to the lack of variation. Compared to NCBI sequences, they showed 97% coverage, an E-value of 0.0, and 97.73% identity with the genus Bassus.
The alignment of the 111 Digonogastra wasps allows us to obtain 563 bp and reveals 7 variable sites. We obtained 100% coverage, 0.0 E-value, and 93.61% identity compared with NCBI sequences. Nucleotide diversity (Pi) for Digonogastra sp. in the peninsula was 0.00228, and haplotype diversity (Hd) was 0.775. The 7 variable sites defined 11 haplotypes. Haplotype 2 was most abundant, followed by haplotypes 4, 7, and 1 (Fig. 1B). Six of the 11 haplotypes were found throughout the entire geographic range, 3 were unique to the northern region, and 2 were only found in the central part of the peninsula. The phylogenetic analysis revealed that 11 haplotypes from the peninsula formed a single clade with 100% support (Fig. S1). This clade was separated from 5 clades found in Canada, although with low bootstrap support (43%).
Figure 3. Mean and standard deviation of 5 significantly different traits of Digonogastra sp. among Baja California Peninsula ecoregions. From north to south, they are listed as follows: Sierra Juárez (SJ), Sierra San Pedro Mártir (SSPM), Chaparral (Ch), Central Desert (DC), Magdalena Plains (PM), and Cape Region (CR). Individuals with larger sizes are marked in red, while those with smaller sizes are marked in blue.
The geographical distance did not cause genetic differentiation in Digonogastra sp. individuals (Fst = -0.01319, S.S. = 41.176, p = 0.84360 ± 0.01326). Similarly, no differentiation was found between individuals inhabiting different ecoregions (Fst = -0.01838, S.S. = 38.933, p = 0.05181 ± 0.00599) and individuals parasitizing different host species (Fst = 0.01877, S.S. = 41.853, p = 0.06158 ± 0.00750). Considering that Digonogastra sp. individuals from the peninsula form a single genetic clade, we performed a demographic analysis (i.e., mismatch analysis), including all individuals as a single population. The distribution of paired differences was unimodal, and the SSD test did not reject the expansion hypothesis (p = 0.09).
Phenotypic variation. Digonogastra sp. females had an average body length of 9.03 mm, with their morphological characters showing coefficients of variation between 10% and 27%. Males had an average body length of 7.34 mm, with morphological characters exhibiting coefficients of variation between 17% and 33%. The most variable character was the metasoma width for females and males, with coefficients of variation of 26.32% and 33.15%, respectively. Females exhibited significant correlations between all analyzed traits (r² > 0.6; p < 0.05), except for ovipositor width, ovipositor tip length, and scape width (r² < 0.3; p > 0.05), while males showed significant correlations across all traits (Fig. S2).
Morphological differences between males and females were significant (F test = 5.21, F = 25.74, p < 0.0001), with females being consistently larger in all measured characters (Table 1). Similarly, significant differences were found among individuals from different ecoregions (Wilks’ Lambda = 0.195, F = 1.83, p < 0.0002; Table 1, Fig. 3), with 11 out of 18 characters showing high variation (Table 1). Finally, significant differences observed in the MANOVA (F test = 0.40, F = 1.98, p < 0.027) indicate that the traits of the wasps vary according to the host groups, even though the ANOVA did not detect significant differences in individual traits (Table 1).
Discussion
The taxonomic, genetic, and phenotypic diversity of parasitoid wasps is influenced by environmental and spatial distribution of their populations and hosts (Althoff & Thompson, 2001; Baer et al., 2004; Harrison et al., 2022). Here, we have recorded for the first time Bassus wasps attacking Tegeticula moths. Additionally, we include the first report of a Digonogastra wasps attacking T. baja. The 111 Digonogastra sp. individuals from various environments showed low genetic diversity across the peninsula, suggesting a single panmictic population that had experienced historical demographic expansion. We also identified sexual dimorphism and morphometric variation due to ecoregions and diferent host.
Genetic diversity estimates for Digonogastra sp. in the Baja California Peninsula are moderate (N = 111, 11 haplotypes, Hd = 0.77, pi = 0.00228). Similar genetic diversity values have been reported for other parasitoid wasps in the Braconidae family (Baer et al., 2004; Hufbauer et al., 2004). For Digonogastra sp., we obtained values of reduced nucleotide diversity alongside high haplotype diversity, indicating that population haplotypes are very similar to each other, as shown in our haplotype network (Fig. 1B). This pattern is observed in species that have experienced population expansion events (Roderick, 1996). Individuals from the Baja California Peninsula have different haplotypes than those recorded to the north of the genus’s distribution, and they cluster into a different clade from the haplotypes reported in Canada (Fig. S1). Future sampling in intermediate areas will help determine whether the diversity found in this study is shared with other zones of their distribution or is restricted to this geographic region.
We found no genetic diversity structuring Digo-
nogastra sp. individuals from different geographic areas, ecoregions, or hosts (i.e., Tegeticula spp.) in the Baja California Peninsula. Furthermore, Digonogastra individuals share most of the recorded haplotypes, suggesting a single panmictic population. A similar pattern was observed in the parasitoid wasp Eusandalum sp., which attacked 11 species of Prodoxus spp. moths in a Yucca complex, in the USA (Althoff, 2008). These 2 wasp genera are known for their generalist nature, which may be related to the genetic differentiation pattern across the landscape. Eusandalum sp., in particular, can lay eggs throughout the year and parasitize any available Prodoxus species, which helps to maintain a continuous population across the landscape (Althoff, 2008). Digonogastra has been recorded attacking various Tegeticula and Prodoxus moths (Force & Thompson, 1984), also present in the peninsula (obs. pers; Althoff et al., 2007). Therefore, other potential food sources could contribute to the population connectivity of Digonogastra sp. throughout its distribution. For example, Prodoxus larvae have been observed as a year-round resource (Powell, 1989). However, further studies are needed to confirm the presence of Digonogastra sp. attacking other Tegeticula and Prodoxus species in this region.
The panmictic population of Digonogastra sp. in the Baja California Peninsula exhibited a historical population expansion, as supported by different analyses (Fig. 1C). The close parasitoid-host interaction with Yucca-pollinating moths implies that demographic changes in their hosts (moths) and in the plants can directly affect the demographics of their populations. Previous studies have recorded the influence of glacial and interglacial cycles in the Quaternary on the demographic history of organisms in the Baja California Peninsula (Garrick et al., 2009; Harrington et al., 2018; Nason et al., 2002). Demographic changes have been documented for the 3 Yucca species in this region, and their habitat has expanded since the last interglacial maximum (Alemán et al., 2024; Arteaga et al., 2020; De la Rosa et al., 2020). Since Digonogastra sp. individuals rely on Yucca plant fruits to complete their life cycle, as these fruits host the moth larvae that serve as their food source, the population expansion found in these wasps may be a consequence of the population expansion observed in the plants that host their hosts.
Phenotypic variability and sexual dimorphism of Digonogastra sp. Geographic variation in phenotype is a common factor in insect populations (Stilwell & Fox, 2007, 2009). The spatial structure of this variation can be determined by environmental conditions, genetic composition, and/or ecological interactions (Agrawal, 2001; Resh & Cardé, 2009; Seifert et al., 2022). For Digonogastra sp. in the Baja California Peninsula, our results show phenotypic variability and a high phenotypic correlation among the studied traits (Table 1, Fig. S2). However, the female ovipositor showed low variation and correlation with the other traits, indicating that its variation did not depend closely on the expression of other traits. Females use the ovipositor to pierce the fruit pericarp, access the moth larvae, and lay eggs (Crabb & Pellmyr, 2006; Resh & Cardé, 2009; Vilhelmsen et al., 2001). The function of this trait is closely related to its fitness, as the arrangement of wasp eggs near moth larvae inside the fruit determines their survival by allowing access to their food source. This may favor reduced variation in ovipositor size (Mazer & Damuth, 2001; Pigliucci, 2003).
Like other parasitoid wasps, Digonogastra sp. exhibits sexual dimorphism (Hurlbutt, 1987; Quicke, 2015), with females being larger in all the traits assessed compared to males (Table 1). Sexual dimorphism in Hymenoptera is partly attributed to complementary sex determination (CSD), where fertilized eggs develop into females and unfertilized eggs into males (Quicke, 2015; Resh & Cardé, 2009). Studies on parasitoid wasps have shown that females typically allocate more resources to fertilized eggs (females) than unfertilized ones (males; Ellers & Jervis, 2003; Jervis et al., 2008; Quicke, 2015; Resh & Cardé, 2009; Visser, 1994). Therefore, the variation in size can be explained by the interplay between CSD and the differential allocation of resources during oviposition.
The environmental heterogeneity in which these wasps are distributed in the Baja California Peninsula also affects their phenotypic variation (Fig. 3). Similar patterns have been observed in butterflies, where species distributed across a broad environmental range exhibit greater variation in organism size compared to species with a more restricted environmental distribution (Seifert et al., 2022). The relationship between body size and environmental variability is attributed to the significant influence of the environment on the development and growth of holometabolous insects, considering factors such as temperature, humidity, and nutrition (Davidowitz et al., 2004; Stillwell & Fox, 2007; Wonglersak et al., 2020). Digonogastra sp. wasps from the Baja California Peninsula occur in different ecosystems, including mountainous areas, deserts, and lowland forests, with variable climatic conditions. However, this study did not investigate the environmental factors that may affect morphological differentiation, which is a topic for future research.
Ecological interactions between plants, herbivores, and parasitoids are significant drivers of biological diversity in terrestrial ecosystems (Schoonhoven et al., 2005). The phenotypic variation in Digonogastra sp. is influenced by the interaction between the wasp, the ecoregion and the host. Tritrophic interactions have shown that a favorable environment for plant growth leads to better nutrition for herbivorous insects, enhancing the development and performance of parasitoids (Han et al., 2019; Pekas & Wäckers, 2020; Schoonhoven et al., 2005). These “bottom-up” cascades have been studied and have shown that the nutritional quality of the plant and the host insect plays a critical role in parasitoids. For instance, it has been observed that parasitoid wasps have increased fitness when they inhabit more fertile soils (Pekas & Wäckers, 2020; Sarfraz et al., 2009). This suggests that the different trophic levels may influence the phenotypic diversity of this wasp, namely the plant and herbivore.
In conclusion, our study records for the first time the genus of parasitoid wasps Bassus attacking Tegeticula moths and increases the diversity of hosts attacked by Digonogastra wasps. The genetic diversity of Digonogastra sp. in the Baja California Peninsula is moderate, forming a single panmictic population with signs of historical demographic expansion. Phenotypic variation is influenced by sexual dimorphism, ecoregions, and their host, this highlights the various factors that can shape the phenotype of these parasitoid wasps. The presence of Digonogastra sp. in different ecoregions suggests the influence of ecological interactions on their phenotypic diversity.
Acknowledgements
The authors are grateful to Leonardo de la Rosa, Mario Salazar, José Delgadillo, and Darlene van der Heiden for their help with laboratory analysis, technical support, and assistance in the fieldwork. C.R.A.H thanks the Centro de Investigación Científica y Educación Superior de Ensenada (CICESE) and Universidad Autónoma de Baja California (UABC) for offering academic support. This study was supported financially by Consejo Nacional de Ciencia y Tecnología (Conacyt) (CB-2014-01-238843, infra-2014-1-226339). The Rufford Foundation also provided financial support for a part of this study (RSG 13704-1) and the Jiji Foundation. The authors thank the Associate Editor and the anonymous reviewer for their valuable comments. The authors do not have any conflict of interest to declare.
Haplotypes from this study were deposited in the GenBank with accession numbers PQ252653-PQ252665.
References
Abdala-Roberts, L., Puentes, A., Finke, D. L., Marquis, R. J., Montserrat, M., Poelman, E. H. et al. (2019). Tri-trophic interactions: bridging species, communities and ecosystems. Ecology Letters, 22, 2151–2167. https://doi.org/10.1111/ele.13392
Agrawal, A. A. (2001). Phenotypic plasticity in the interactions and evolution of species. Science, 294, 321–326. https://doi.org/10.1126/science.1060701
Althoff, D. M., & Thompson, J. N. (2001). Geographic structure in the searching behaviour of a specialist parasitoid: combining molecular and behavioural approaches. Journal of Evolutionary Biology, 14, 406–417. https://doi.org/10.1046/j.1420-9101.2001.00286.x
Althoff, D. M., Svensson, G. P., & Pellmyr, O. (2007). The influence of interaction type and feeding location on the phylogeographic structure of the yucca moth community associated with Hesperoyucca whipplei. Molecular Phy-
logenetics and Evolution, 43, 398–406. https://doi.org/10.
1016/j.ympev.2006.10.015
Althoff, D. M. (2008). A test of host-associated differentiation across the ‘parasite continuum’in the tri-trophic interaction among yuccas, bogus yucca moths, and parasitoids. Molecular Ecology, 17, 3917–3927. https://doi.org/10.1111/j.1365-294X.2008.03874.x
Althoff, D. M., Segraves, K. A., Smith, C. I., Leebens-Mack, J., & Pellmyr, O. (2012). Geographic isolation trumps coevolution as a driver of yucca and yucca moth diversification. Molecular Phylogenetics and Evolution, 62, 898–906. https://doi.org/10.1016/j.ympev.2011.11.024
Althoff, D. M., & Segraves, K. A. (2022). Evolution of antag-
onistic and mutualistic traits in the yucca-yucca moth obligate pollination mutualism. Journal of Evolutionary Biology, 35, 100–108. https://doi.org/10.1111/jeb.13967
Arteaga, M. C., Bello-Bedoy, R., & Gasca-Pineda, J. (2020). Hybridization between yuccas from Baja California: Genomic and environmental patterns. Frontiers in Plant Science, 11, 685. https://doi.org/10.3389/fpls.2020.00685
Baer, C. F., Tripp, D. W., Bjorksten, T. A., & Antolin, M. F. (2004). Phylogeography of a parasitoid wasp (Diaeretiella rap-
ae): no evidence of host-associated lineages. Molecular Ecology, 13, 1859–1869. https://doi.org/10.1111/j.1365-294X.
2004.02196.x
Bandelt, H. J., Forster, P., & Röhl, A. (1999). Median-joining networks for inferring intraspecific phylogenies. Molecular Biology and Evolution, 16, 37–48. https://doi.org/10.1093/oxfordjournals.molbev.a026036
Carmona, D., Fitzpatrick, C. R., & Johnson, M. T. (2015). Fifty years of co-evolution and beyond: integrating co-evolution from molecules to species. Molecular Ecology, 24, 5315–5329. https://doi.org/10.1111/mec.13389
Crabb, B. A., & Pellmyr, O. (2006). Impact of the third trophic level in an obligate mutualism: do yucca plants benefit from parasitoids of yucca moths? International Journal of Plant Sciences, 167, 119–124. https://doi.org/10.1086/497844
Cuautle, M., & Rico-Gray, V. (2003). The effect of wasps and ants on the reproductive success of the extrafloral nectaried plant Turnera ulmifolia (Turneraceae). Functional Ecology, 17, 417–423. https://doi.org/10.1046/j.1365-2435.2003.00732.x
Darriba, D., Taboada, G. L., Doallo, R., & Posada, D. (2012). jModelTest 2: more models, new heuristics and parallel computing. Nature Methods, 9, 772. https://doi.org/10.1038/nmeth.2109
Davidowitz, G., D’Amico, L. J., & Nijhout, H. F. (2004). The effects of environmental variation on a mechanism that controls insect body size. Evolutionary Ecology Research, 6, 49–62.
De la Rosa-Conroy, L., Gasca-Pineda, J., Bello-Bedoy, R., Eguiarte, L. E., & Arteaga, M. C. (2020). Genetic patterns and changes in availability of suitable habitat support a colonization history of a North American perennial plant. Plant Biology, 22, 233–242. https://doi.org/10.1111/plb.13053
Ellers, J., & Jervis, M. (2003). Body size and the timing of egg production in parasitoid wasps. Oikos, 102, 164–172. https://doi.org/10.1034/j.1600-0706.2003.12285.x
Engelmann, G. (1872). The flower of yucca and its fertilization. Bulletin of the Torrey Botanical Club, 3, 33–33.
Excoffier, L., Laval, G., & Schneider, S. (2005). Arlequin (version 3.0): an integrated software package for population genetics data analysis. Evolutionary Bioinformatics Online, 2005,47–50. https://doi.org/10.1177/117693430500100003
Folmer, O., Hoeh, W. R., Black, M. B., & Vrijenhoek, R. C. (1994). Conserved primers for PCR amplification of mitochondrial DNA from different invertebrate phyla. Molecular Marine Biology and Biotechnology, 3, 294–299.
Force, D. C., & Thompson, M. L. (1984). Parasitoids of the immature stages of several southwestern yucca moths. The Southwestern Naturalist, 29, 45–56. https://doi.org/
10.2307/3670768
Garrick, R. C., Nason, J. D., Meadows, C. A., & Dyer, R. J. (2009). Not just vicariance: phylogeography of a Sonoran Desert euphorb indicates a major role of range expansion along the Baja peninsula. Molecular Ecology, 18, 1916–1931. https://doi.org/10.1111/j.1365-294X.2009.04148.x
Godfray, H. C. J. (1994). Parasitoids: behavioral and evolut-
ionary ecology. New Jersey: Princeton University Press.
González-Abraham, C. E., Garcillán, P. P., & Ezcurra, E. (2010). Ecorregiones de la península de Baja California: una síntesis. Boletín de la Sociedad Botánica de México, 87, 69–82. https://doi:10.17129/botsci.302
Hall, T. A. (1999). BioEdit: a user-friendly biological sequence alignment editor and analysis program for Windows 95/98/NT. Nucleic Acids Symposium Series, 41,95–98.
Han, P., Desneux, N., Becker, C., Larbat, R., Le Bot, J., Adamowicz, S. et al. (2019). Bottom-up effects of irrigation, fertilization and plant resistance on Tuta absoluta: implications for Integrated Pest Management. Journal of Pest Science, 92, 1359–1370. https://doi.org/10.1007/s10340-018-1066-x
Harrington, S. M., Hollingsworth, B. D., Higham, T. E., & Reeder, T. W. (2018). Pleistocene climatic fluctuations drive isolation and secondary contact in the red diamond rattlesnake (Crotalus ruber) in Baja California. Journal of Biogeography, 45, 64–75. https://doi.org/10.1111/jbi.13114
Harrison, K., Tarone, A. M., DeWitt, T., & Medina, R. F. (2022). Predicting the occurrence of host-associated differentiation in parasitic arthropods: a quantitative literature review. Entomologia Experimentalis et Applicata, 170, 5–22. https://doi.org/10.1111/eea.13123
Hebert, P. D., Ratnasingham, S., Zakharov, E. V., Telfer, A. C., Levesque-Beaudin, V., Milton, M. A. et al. (2016). Counting animal species with DNA barcodes: Canadian insects. Philosophical Transactions of the Royal Society B: Biological Sciences, 371, 20150333. https://doi.org/10.1098/rstb.2015.0333
Heil, M. (2008). Indirect defence via tritrophic interactions.
New Phytologist, 178, 41–61. https://doi.org/10.1111/j.1469-
8137.2007.02330.x
Hufbauer, R. A., Bogdanowicz, S. M., & Harrison, R. G. (2004). The population genetics of a biological control introduction: mitochondrial DNA and microsatellie variation in native and introduced populations of Aphidus ervi, a parisitoid wasp. Molecular Ecology, 13, 337–348. https://doi.org/10.
1046/j.1365-294X.2003.02084.x
Hurlbutt, B. (1987). Sexual size dimorphism in parasitoid wasps. Biological Journal of the Linnean Society, 30, 63–89. https://doi.org/10.1111/j.1095-8312.1987.tb00290.x
Jervis, M. A., Ellers, J., & Harvey, J. A. (2008). Resource acquisition, allocation, and utilization in parasitoid reproductive strategies. Annual Review of Entomology, 53,361–385. https://doi.org/10.1146/annurev.ento.53.103106.093
433
Kankare, M., Van Nouhuys, S., & Hanski, I. (2005). Genetic divergence among host-specific cryptic species in Cotesia melitaearum aggregate (Hymenoptera: Braconidae), parasitoids of checkerspot butterflies. Annals of the Entomological Society of America, 98, 382–394. https://doi.org/10.1603/0013-8746(2005)098[0382:GDAHCS]2.0.CO;2
Kappers, I. F., Hoogerbrugge, H., Bouwmeester, H. J., & Dicke, M. (2011). Variation in herbivory-induced volatiles among cucumber (Cucumis sativus L.) varieties has consequences for the attraction of carnivorous natural enemies. Journal of Chemical Ecology, 37, 150–160. https://doi.org/10.1007/s10886-011-9906-7
Karns, G. (2009). Genetic differentiation of the parasitoid, Cotesia congregata (Say), based on host-plant complex (M. Sc. Thesis). Virginia Commonwealth University. VA, USA. https://doi.org/10.25772/1E5V-N037
Lenz, L. W. (1998). Yucca capensis (Agavaceae, Yuccoideae), a new species from Baja California Sur, Mexico. Cactus and Succulent Journal, 70, 289–296.
Lozier, J. D., Roderick, G. K., & Mills, N. J. (2009). Molecular markers reveal strong geographic, but not host associated, genetic differentiation in Aphidius transcaspicus, a parasitoid of the aphid genus Hyalopterus. Bulletin of Entomological Research, 99, 83–96. https://doi.org/10.1017/S0007485308006147
Mazer, S. J., & Damuth, J. (2001). Nature and causes of variation. In C. W. Fox, D. A. Roff, & D. J. Fairbairn (Ed). Evolutionary ecology: concepts and case studies (pp. 3–15). Oxford, UK: Oxford University Press.
Nason, J. D., Hamrick, J. L., & Fleming, T. H. (2002). Historical vicariance and postglacial colonization effects on the evolution of genetic structure in Lophocereus, a Sonoran Desert columnar cactus. Evolution, 56, 2214–2226. https://doi.org/10.1111/j.0014-3820.2002.tb00146.x
Navascués, M., Vaxevanidou, Z., González-Martínez, S. C., Climent, J., Gil, L., & Emerson, B. C. (2006). Chloroplast microsatellites reveal colonization and meta-
population dynamics in the Canary Island pine. Molecular Ecology, 15, 2691–2698. https://doi.org/10.1111/
j.1365-294X.2006.02960.x
Nei, M., & Li, W. H. (1979). Mathematical model for studying genetic variation in terms of restriction endonucleases. Proceedings of the National Academy of Sciences, 76, 5269–5273. https://doi.org/10.1073/pnas.76.10.5269
Nei, M. (1987). Molecular evolutionary genetics. New York: Columbia University Press.
Pekas, A., & Wäckers, F. L. (2020). Bottom-up effects on tri-trophic interactions: Plant fertilization enhances the fitness of a primary parasitoid mediated by its herbivore host. Journal of Economic Entomology, 113, 2619–2626. https://doi.org/10.1093/jee/toaa204
Pellmyr, O., & Leebens-Mack, J. (1999). Forty million years of mutualism: evidence for Eocene origin of the yucca-yucca moth association. Proceedings of the National Academy of Sciences, 96, 9178–9183. https://doi.org/10.1073/pnas.96.
16.9178
Pellmyr, O. (2003). Yuccas, yucca moths, and coevolution: a review. Annals of the Missouri Botanical Garden, 90, 35–55. https://doi.org/10.2307/3298524
Pigliucci, M. (2003). Phenotypic integration: studying the ecology and evolution of complex phenotypes. Ecology Letters, 6, 265–272. https://doi.org/10.1046/j.1461-0248.2003.00428.x
Powell, J. A. (1989). Synchronized, mass-emergences of a yucca moth, Prodoxus Y-inversus (Lepidoptera: Prodoxidae), after 16 and 17 years in diapause. Oecologia, 81, 490–493. https://doi.org/10.1007/BF00378957
Quicke, D. L. (2015). The Braconid and Ichneumonid parasitoid wasps: Biology, Systematics, Evolution and Ecology. Metopiinae. Oxford: Wiley Blackwell.
Resh, V. H., & Cardé, R. T. (Eds.). (2009). Encyclopedia of insects. San Diego, CA: Academic press.
Roderick, G. K. (1996). Geographic structure of insect populations: gene flow, phylogeography, and their uses. Annual Review of Entomology, 41, 325–352. https://doi.org/10.1146/annurev.en.41.010196.001545
Rogers, A. R., & Harpending, H. (1992). Population growth makes waves in the distribution of pairwise genetic differences. Molecular Biology and Evolution, 9, 552–569. https://doi.org/10.1093/oxfordjournals.molbev.a040727
Rozas, J., Sánchez-DelBarrio, J. C., Messeguer, X., & Rozas, R. (2003). DnaSP, DNA polymorphism analyses by the coalescent and other methods. Bioinformatics, 19, 2496–2497. https://doi.org/10.1093/bioinformatics/btg359
Sánchez, J. A., Romero, J., Ramírez, S., Anaya, S., & Carrillo, J. L. (1998). Géneros de Braconidae del estado de Guanajuato (Insecta: Hymenoptera). Acta Zoológica Mexicana (nueva serie), 79, 59–137. https://doi.org/10.21829/azm.1998.74741721
Sarfraz, M., Dosdall, L. M., & Keddie, B. A. (2009). Host plant nutritional quality affects the performance of the parasitoid Diadegma insulare. Biological Control, 51, 34–41. https://doi.org/10.1016/j.biocontrol.2009.07.004
Schoonhoven, L. M., Van Loon, J. J., & Dicke, M. (2005). Insect-plant biology. Oxford: Oxford University Press.
Seifert, C. L., Strutzenberger, P., & Fiedler, K. (2022). Ecological specialisation and range size determine intraspecific body size variation in a speciose clade of insect herbivores. Oikos, 2022, e09338. https://doi.org/10.1111/oik.09338
Singer, M. S., & Stireman III, J. O. (2005). The tri-trophic niche concept and adaptive radiation of phytophagous insects. Ecology Letters, 8, 1247–1255. https://doi.org/10.
1111/j.1461-0248.2005.00835.x
Stillwell, R. C., & Fox, C. W. (2007). Environmental effects on sexual size dimorphism of a seed-feeding beetle. Oecologia, 153, 273–280. https://doi.org/10.1007/s00442-007-0724-0
Stillwell, R. C., & Fox, C. W. (2009). Geographic variation in body size, sexual size dimorphism and fitness components of a seed beetle: local adaptation versus phenotypic plasticity. Oikos, 118, 703–712. https://doi.org/10.1111/j.1600-0706.2008.17327.x
StiremanIII, J. O., Nason, J. D., & Heard, S. B. (2005). Host-associated genetic differentiation in phytophagous insects: general phenomenon or isolated exceptions? Evidence from a goldenrod-insect community. Evolution, 59, 2573–2587. https://doi.org/10.1111/j.0014-3820.2005.tb00970.x
Takabayashi, J., & Dicke, M. (1996). Plant-carnivore mutualism through herbivore-induced carnivore attractants. Trends
in plant science, 1, 109–113. https://doi.org/10.1016/S1360-
1385(96)90004-7
Tamura, K., Dudley, J., Nei, M., & Kumar, S. (2007). MEGA4: molecular evolutionary genetics analysis (MEGA) software version 4.0. Molecular Biology and Evolution, 24, 1596–1599. https://doi.org/10.1093/molbev/msm092
Turner, R. M., Bowers, J. E., & Brugess, T. L. (2022). Sonoran Desert plants: an ecological atlas. Tucson: University of Arizona Press.
Vilhelmsen, L., Isidoro, N., Romani, R., Basibuyuk, H. H., & Quicke, D. L. (2001). Host location and oviposition in a basal group of parasitic wasps: the subgenual organ, ovipositor apparatus and associated structures in the Orussidae (Hymenoptera, Insecta). Zoomorphology, 121, 63–84. https://doi.org/10.1007/s004350100046
Visser, M. E. (1994). The importance of being large: the relationship between size and fitness in females of the parasitoid Aphaereta minuta (Hymenoptera: Braconidae). Journal of Animal Ecology, 63, 963–978. https://doi.org/10.
2307/5273
Wonglersak, R., Fenberg, P. B., Langdon, P. G., Brooks, S. J., & Price, B. W. (2020). Temperature-body size responses in insects: a case study of British Odonata. Ecological Entomology, 45, 795–805. https://doi.org/10.1111/een.12853
Diversity of anurans and use of microhabitatsin three vegetation coverages of the Santuario de Flora y Fauna Los Colorados, Colombian Caribbean
Omer José Jiménez-Ortega a, d, Keiner L. Tílvez b, Joselin Castro-Palacios a,
Andrés García c, *, Gabriel R. Navas a, Julio Abad Ferrer-Sotelo e, Dilia Naranjo-Calderón e, Juan Gabriel Díaz-Castellar e, Víctor Buelvas-Meléndez e
a Universidad de Cartagena, Campus San Pablo, Grupo de Investigación en Hidrobiología, Programa de Biología, Carrera 50#24-120, Zaragocilla, Cartagena de Indias, Colombia.
b Universidad de Cartagena, Campus San Pablo, Grupo de Investigación en Biología Descriptiva y Aplicada, Carrera 50#24-120, Zaragocilla, Cartagena de Indias, Colombia
c Universidad Nacional Autónoma de México, Instituto de Biología, Estación de Biología Chamela, Apartado postal 21, 48980 San Patricio, Jalisco, México
d Parque Temático Vivarium del Caribe-Fundación Archosauria zona norte km 15, Provincia de Cartagena, Bolívar, Colombia
e Santuario de Flora y Fauna Los Colorados, Parques Nacionales Naturales de Colombia, Carrera 8# 9-20 Plaza Olaya Herrera, San Juan Nepomuceno, Bolívar, Colombia
*Corresponding author: chanoc@ib.unam.mx (A. García)
Received: 28 October 2023; accepted: 18 March 2024
Abstract
This study aimed to determine anuran diversity and the use of microhabitats in 3 vegetation covers in the Santuario de Flora y Fauna Los Colorados. Five field trips of 6 days each were made, 2 days and 2 nights in each cover: forest, pasture, and crop. Sampling was carried out with the visual encounter inspection technique under a randomized design by random walks with manual capture. A total of 19 species were recorded, 14 in the forest, 13 in pasture, and 12 in crop. Pasture and crop were the vegetation covers with the greatest similarity of species. This work updates the list of anuran species recorded in the management plan of the Santuario de Flora y Fauna Los Colorados 2018-2023. The greatest number of anuran species was associated with leaf litter, “jagüeyes”, and soils. The transformation of the landscape as a result of agriculture and cattle ranching generated changes in the richness, abundance, composition, and use of microhabitats of the anurans present in the Santuario de Flora y Fauna Los Colorados.
Keywords: Landscape transformation; Vegetation coverage; Microhabitat; Tropical dry forest
Diversidad de anuros y uso de microhábitats en tres coberturas vegetales del Santuario de Flora y Fauna Los Colorados, Caribe colombiano
Resumen
Este estudio tuvo como objetivo determinar la diversidad de anuros y el uso de microhábitats en 3 coberturas vegetales en el Santuario de Flora y Fauna Los Colorados. Se hicieron 5 salidas de campo de 6 días cada una, 2 días y 2 noches en cada una: bosque, potrero y cultivo. Se realizaron muestreos con la técnica de inspección por encuentro visual, bajo el diseño aleatorizado por caminatas al azar con captura manual. Se registraron 19 especies, 14 de ellas en bosque, 13 en potrero y 12 en cultivo, siendo el potrero y el cultivo las coberturas con mayor similitud de especies. Este trabajo actualiza el listado de las especies de anuros registrados en el Plan de manejo del Santuario de Flora y Fauna Los Colorados 2018-2023. El mayor número de especies de anuros se encontró asociado a la hojarasca, el jagüey y los suelos. La transformación del paisaje producto de la agricultura y la ganadería genera cambios en la riqueza, abundancia, composición y uso de microhábitats de los anuros presentes en el Santuario de Flora y Fauna Los Colorados.
Palabras clave: Transformación del paisaje; Coberturas vegetales; Microhábitat; Bosque seco tropical
© 2024 Universidad Nacional Autónoma de México, Instituto de Biología. This is an open access article under the CC BY-NC-ND license
(http://creativecommons.org/licenses/by-nc-nd/4.0/)
Introduction
Seasonally tropical dry forests (STDF here after) in Colombia are distributed mainly in the inter-Andean valleys and the Caribbean region (García et al., 2014), the latter being one of the regions with the best conserved areas of this ecosystem (Pizano et al., 2014; Rodríguez et al., 2012). However, Etter et al. (2008) point out that indiscriminate deforestation for various anthropogenic activities such as agriculture and livestock have generated large reductions in forest cover over time. This, combined with other activities such as mining and urban development (Cristal et al., 2020; Galván-Guevara et al., 2015; Jiménez et al., 2018), cause biological and ecological interactions to deteriorate, and the functionality of the ecosystem is compromised (Thomson et al., 2017), which is why Colombian STDFs have been classified as critically endangered (CR) (Etter et al., 2017). Consequently, it is a strategic ecosystem for conservation study due to its high risk of disappearing, strongly threatening the local fauna and the people who depend directly and indirectly on the ecosystem services it provides (Andrade, 2011).
One of most sensitive groups to forest transformation is amphibians, including anurans, which are highly dependent on humid places or sites with high water availability since most of their species have indirect development, permeable skin, and anamniote-type eggs (O’Malley, 2007). The spatial distribution and microhabitats use by anurans depend on the physiological requirements of each organism, and the available resources (Urbina-Cardona et al., 2006; Zug et al., 2009), as suggested by several studies showing many anuran species prefer forested areas (Cáceres-Andrade & Urbina-Cardona, 2009; García-R et al., 2005; Román-Palacios et al., 2016). Consequently, these species may be affected by anthropogenic disturbance, forest fragmentation, and loss (Cáceres-Andrade & Urbina-Cardona, 2009).
Forest transformation is among the main factors affecting anuran communities (Cáceres-Andrade & Urbina-Cardona, 2009; Marín et al., 2017; Romero, 2013; Vargas & Bolaños-L, 1999), causing around 38% of Colombian amphibians to be included under a category of endangered species and positioning Colombia as the country with the highest number of threatened species according to the second global review of amphibians (Re:wild, 2023). A study carried out by Duarte-Marín et al. (2018) in 3 habitats of the Selva de Florencia National Natural Park estimated that the covers with greater vegetation (forest and pine forest) presented greater richness and diversity of anurans than those covers with less vegetal complexity (pastures). This means land use and changes in vegetation cover are factors that influence amphibian species richness and diversity. Therefore, species that are not adapted to the new environmental conditions created by landscape transformation are eliminated from the assembly, negatively affecting the ecosystem processes in which they had intervention (Díaz et al., 2006).
Additionally, forest fragmentation has created barriers that prevent anuran dispersal, resulting in a decrease in their genetic diversity (De Sá, 2005). Furthermore, it has generated changes in the composition and abundance of anurans to an extent that depends on the levels of disturbance (Acuña-Vargas, 2016), with an increase in the penetration of light and winds along the perimeter of a forest remnants, coming from non-forest environments such as pastures, with the subsequent changes in microclimates (Echeverry et al., 2006; Galván-Guevara et al., 2015; Laurence & Gascon, 1997), phenomena known as the edge effect ( Rojas & Pérez-Peña, 2018). However, studies such as Blanco and Bonilla (2010) show that some transformed areas provide a greater number of microenvironments due to the modifications made by humans (e.g., creation of jagüeyes) and record greater richness and abundance of anurans species when compared to less transformed areas, which is known as intermediate disturbance theory (Conell, 1978). However, it must be considered that the species found in these areas have extensive plasticity to tolerate the environmental and structural gradients generated by anthropogenic disturbance, that is, they are resilient (Cáceres-Andrade & Urbina-Cardona, 2009).
Based on the above, the general objective of this research was to determine the diversity of anurans and their use of microhabitats in 3 vegetation covers within the Los Colorados Flora and Fauna Sanctuary (SFF Los Colorados), an important protected area in the Caribbean region of Colombia, which contributes to the understanding of how amphibians respond to changes in land use for agriculture and livestock, in order to provide information that can be useful for environmental entities to determine management and conservation policies for these organisms in landscape fragments.
The specific objectives are: 1) to determine the richness, abundance, diversity, and composition of anurans in 3 vegetation covers that are representative of the Los Colorados SFF; 2) to describe the use of the microhabitat by the species in each vegetation cover; 3) to analyze and compare the relationship between precipitation and environmental temperature with the richness, abundance, and diversity of species in each vegetation cover and; 4) to analyze and compare the alpha and beta diversity of anuran species in each vegetation cover.
We expect to record differences between the 3 types of vegetation cover, hypothesizing that due to the greater heterogeneity of an ecosystem in better conservation condition such as the tropical forest, it will register a greater richness and diversity of species and its species composition will differ with respect to the other covers. While the use of the microhabitat by the species will differ in each cover and will depend on the variety of natural or anthropogenic substrates existing in each site.
Materials and methods
Montes de María is a subregion of the Colombian Caribbean. It is located between the departments of Sucre and Bolívar with an area of 6,297 km2, of which 3,719 km2 belong to the department of Bolívar (Aguilera-Díaz, 2013; Herazo et al., 2017). It integrates several municipalities, among which is San Juan Nepomuceno, where the SFF Los Colorados is located at 9°56’06.7” N, 75°06’48.7” W (Fig. 1) with an area of 1,041.96 ha, an average high temperature of 28 °C and an elevation of 23 m asl (Jiménez et al., 2018). Due to the seasonality of rainfall in the region, 3 seasons can be identified, each lasting 4 months and including the dry season (December to April), the transition season (little rain, May to August), and the rainy season (abundant rain, September to November). The average precipitation is around 1,643mm with a monthly average of 137mm (Rangel & Carvajal-Cogollo, 2012).
SFF Los Colorados is composed of a small mountain system formed by sedimentary rocks, in which the largest and most important STDF relic of Montes de María is located (Jiménez et al., 2018). This ecosystem has humid forest components, which is why it is considered a place of high species diversity (IAVH, 1998). Its hydrographic system is made up of 2 streams: Cacaos and Salvador, located on the south and north sides, respectively; it also has a large number of ravines that flow into these streams (Jiménez et al., 2018). There are 6 land uses within the SFF Los Colorados (Fig. 1), which are in descending order by their percentage of coverage, forest (66.36%), agricultural areas (17.29%), pastures (12.01%), herbaceous and shrubby vegetation areas (3.51%), urbanized areas (0.80%), and mining extraction areas (0.02%). The exact age of the crop areas is unknown; this area has historically been agricultural, even before 1977 when the SFF Los Colorados was declared as a protected natural area. However, for about 10 years these areas have been in the succession stage towards shrublands because they were purchased and practically little cultivated. There are only crops at the sampling point where yam (Dioscorea) or tuber is grown. The only management that is done with these crops is slash-and-burn. With respect to livestock, none of the pasture areas in the sampling sites have more than 40 heads of livestock. No fertilizers or other types of agrochemicals or pesticides are used.
SFF Los Colorados faces 2 main problems in the conservation of their natural environments. The first is an occupancy rate close to 30% of its surface (3 neighborhoods and 11 properties). The second is the inadequate environmental planning outside the protected area that has generated a transformation of the landscape because of cattle ranching, agriculture, forest plantations, mining activities, the presence of a national highway as a limit, and the proximity to a municipal seat of 25,000 inhabitants (Jiménez et al., 2018).
A two-day prospecting visit was carried out at the SFF Los Colorados in November 2021 to inspect the site and locate the sampling points. Subsequently, 5 field trips of 6 days each were carried out (2 days and 2 nights in each cover: forest, pasture, and crop) during the months of January, February, March, April, and June 2022. In this way, sampling was carried out during the dry and transition season, that is, under conditions of no rain (February to April) or very little rain (June). In these months 2 researchers and 2 officials from the SFF Los Colorados carried out daytime (8:00 -10:00 am) and nighttime (6:00-8:00 pm) outings with a constant speed route, for a sampling effort of 160 man-hours in each coverage for a total of 480 man-hours.
The visual encounter inspection technique was used to locate and record anuran species and their abundance, under the randomized design of random walks (Crump & Scott, 2001) and manual capture of individuals (Aguirre-León, 2011; Manzanilla & Péfaur, 2000). The identification of anuran species in each cover (forest, pasture, and crop) was based on regional taxonomic keys (Ballesteros-Correa et al., 2019; Cuentas et al., 2002; Dunn, 1994), supported by field guides with photographs (Meza-Tílvez et al., 2018; Salvador & Gómez-Sánchez, 2018), and databases (Acosta-Galvis, 2021).
The 3 selected coverages were described following the CORINE Land Cover methodology adapted for Colombia (IDEAM et al., 2008) as follows: forest is an area made up mainly of tree elements of native or exotic species, trees being woody plants with a single main trunk or in some cases with several stems, which also have a defined and semi continuous canopy. In the study area, trees reach a height greater than 5m, and watercourses with a width of less than 50m were found. Pasture includes lowlands covered with grasses and some scattered trees with a height greater than 5 m, which are located on hills and flat pastures in warm climates. Crops are areas dedicated primarily to the production of food, fiber, and other raw materials with permanent, transitional, or annual crops of avocado, chili and cassava. Temporary yam crops are mainly found in the study area.
To describe microhabitat used by anurans, the number of individuals of each species observed in one of the substrate types (leaf litter, branch, trunk, sites with the presence of water, rock, soil, herbaceous or shrubby vegetation) were recorded (Cáceres-Andrade & Urbina-Cardona, 2009).
Figure 1. Location of the Los Colorados Flora and Fauna Sanctuary; source National Natural Parks of Colombia, with permission granted by SFF Los Colorados.
All observed species were photographed and at least 1 individual per species was collected, anesthetized with 2% xylocaine gel on the head and belly, and sacrificed (McDiarmid, 1994). To avoid tissue necrosis, they were prepared and fixed with 10% formalin (McDiarmid, 1994; Simmons & Muñoz-Saba, 2005), then placed in a suitable position in a container that had a white absorbent paper impregnated with 10% formalin. Distinctive characteristics were then observed. Finally, they were preserved in 70% ethanol (Cortez-F et al., 2006). The collected material was deposited in the Armando Dugand Gnecco collection of the Universidad del Atlántico, with the following catalog numbers: UARC-Am-00508, UARC-Am-00509, UARC-Am-00510, UARC-Am-00511, UARC- Am-00512, UARC-Am-00513, and UARC-Am-00514. The collecting permit was granted by the regional environmental authority called the Regional Autonomous Corporation of the Canal del Dique (Cardique), and the permit number is the resolution number 0751 of June 27, 2014. In addition, through the research endorsement approved by National Parks of Colombia, No. 20212000004933, October 25, 2021.
Information on the number of species and their abundance in each cover and climatic season was stored in an Excel. To confirm sampling was carried out on dates with the typical characteristics of the climatic seasons (rainy and dry), we graphed and compared statistically (ANOVA) the average precipitation and temperature for the months in which the sampling was carried out based on data obtained from the Institute of Hydrology, Meteorology and Environmental Studies (IDEAM) of the Guamo-Bolívar Station (Retrieved on July 19, 2022, from: http://www.ideam.gov.co/web/atencion-y-participacion-ciudadana/pqrs).
To detect significant differences in alpha diversity (richness, abundance, Simpson, Shannon), the Kruskall-Wallis or ANOVA tests were applied, depending on the normality of the data using the Shapiro-Wilk test and homogeneity of variances using Levene’s test (p < 0.05).
Alpha diversity was determined as the species richness for each coverage (Moreno, 2001), and was evaluated using Chao 1, 2, and Jack 1 estimators in EstimateS v. 9.1 (Villareal et al., 2004). In addition, bootstrap was used, which is useful to determine richness with a high number of rare species (Colwell & Coddington, 1994; Magurran, 2004). On the other hand, the diversity of anurans was estimated for each cover using the Shannon-Wiener index in the program PAST v. 4.03 (Hammer et al., 2001), and dominance using the Simpson index, where values close to 0 were considered as low levels of dominance and those close to 1 as high levels of dominance (Clarke et al., 2014).
To evaluate the turnover of anuran species between different covers (forest, pasture, and crops), the Jaccard index was used because it relates the number of shared species to the total number of exclusive species (Villareal et al., 2004). The range of values goes from 0 in the case of no shared species, to 1 when the covers have the same species composition (Moreno, 2001). From the estimator, a dendrogram was constructed in PAST v. 4.03.
To analyze the use of microhabitats, a graph was constructed where the percentage of use of each microhabitat by species and cover was established, to observe in each cover which microhabitats were most used by each species of anuran. Data were plotted in Excel.
Results
In total 1,269 individuals belonging to 19 species and 1 casual record (not included in this analysis) were recorded and grouped into 13 genera and 7 families (Table 1). Hylidae and Leptodactylidae were the families with the greatest species recorded, 8 and 6, respectively whereas only 1 species was recorded for Microhylidae and Phyllomedusidae.
Table 1
Taxonomic list and number of anuran individuals recorded in forest, crop, and pasture cover in the Los Colorados Flora and Fauna Sanctuary.
Family | Species | Forest | Crops | Pasture |
Bufonidae | Rhinella horribilis (Wiegmann, 1833) | 29 | 48 | 89 |
Rhinella humboldti (Spix, 1824) | 2 | 97 | 101 | |
Ceratophryidae | Ceratophrys calcarata (Boulenger, 1890)* | |||
Dendrobatidae | Dendrobates truncatus (Cope, 1861, “1860”) | 173 | ||
Hylidae | Boana platanera (Escalona et al., 2021) | 23 | 3 | 4 |
Boana pugnax (Schmidt, 1857) | 3 | 90 | ||
Dendropsophus ebraccatus (Cope, 1874) | 1 | |||
Dendropsophus microcephalus (Cope, 1886) | 2 | 4 | 59 | |
Scarthyla vigilans (Solano, 1971) | 3 | |||
Scinax cf. rostratus (Peters, 1863) | 5 | 14 | ||
Scinax cf. ruber (Laurenti, 1768) | 1 | 2 | ||
Trachycephalus typhonius (Linnaeus, 1978) | 6 | 1 | 7 | |
Leptodactylidae | Engystomops pustulosus (Cope, 1864) | 190 | 15 | 12 |
Leptodactylus fuscus (Schneider, 1799) | 2 | 55 | ||
Leptodactylus insularum (Barbour, 1906) | 12 | 1 | 81 | |
Leptodactylus poecilochilus (Cope, 1862) | 44 | |||
Leptodactylus savagei (Heyer, 2005) | 17 | |||
Pleurodema brachyops (Cope, 1869, “1868”) | 38 | 10 | ||
Microhylidae | Elachistocleis panamensis (Dunn et al., 1948) | 19 | ||
Phyllomedusidae | Phyllomedusa venusta (Duellman & Trueb, 1967) | 1 | 5 | |
* Species recorded casually outside the sampled coverage, which is not included in the analyses of our study. |
Table 2
Richness estimators and percentages of representativeness with respect to the number of anuran species recorded in the 3 coverages of the SFF Los Colorados.
Richness estimator | Cover | ||
Forest | Pasture | Crops | |
Species recorded | 14 | 13 | 12 |
Chao 1 | 15.00 (86.7%) | 13.00 (100.0%) | 12.33 (97.3%) |
Chao 2 | 14.68 (88.6%) | 13.00 (100.0%) | 12.90 (93.0%) |
Jack 1 | 16.70 (77.8%) | 13.90 (93.5%) | 14.70 (81.6%) |
Bootstrap | 15.40 (84.3%) | 13.57 (95.8%) | 13.32 (90.1%) |
Alpha diversity was highest in forest (14 species), followed by pasture (13 species) and crop (12 species). The species accumulation curves in the 3 coverages based on the Chao 1, Chao 2, and bootstrap estimators allowed estimating a number of species similar to that recorded in the field and an efficiency in the sampling carried out with a representativeness greater than 80%. The Jack 1 estimator for forest indicates a representativeness of 77.8%, and for pasture and crops greater than 80% (Table 2, Fig. 2). In the singleton and doubleton curves (Fig. 2), a decreasing behavior is observed for the pasture and crop covers, indicating little probability of finding new anuran species in these covers. For forest, the doubleton curve shows an ascending behavior, indicating a probability of finding more species in this cover.
Figure 2. Accumulation curves of anuran species for the 3 coverages of the Los Colorados Flora and Fauna Sanctuary.
Figure 3. Box plots of richness (A), abundance (B), Shannon Index (C), and Simpson (D) for each of the 3 vegetation covers.
Figure 4. Histogram and box plot of daily ambient temperature across the sampling months; temperature (A, B), rainfall (C).
According to the determined Shannon-Wiener index, the diversity for forest cover was 1,631, crops 1,677, and pasture 2,107. On the other hand, Simpson’s index estimated a dominance of 0.726 for the forest, crops 0.746, and pasture 0.858. When comparing the metrics recorded in each vegetation cover (Fig. 3), the pastures registered on average the greatest richness, abundance, and diversity. The average richness was similar between the crops and the forest; however, the variation was greater in the crops. In contrast, the average and variation of abundance was greater in the forest than in the crops. The average species diversity was lowest in forests, intermediate in crops, and highest in pastures. There were statically differences of all metrics among vegetation cover; richness (one-way ANOVA, F = 5.456, df = 2, p > 0.05), abundance (H(χ2) = 4.63, p > 0.05), Shannon (one-way ANOVA, F = 16.71, df = 2, p > 0.05), and Simpson (H(χ2) = 15.97, p > 0.05).
When we graph the monthly fluctuations of ambient temperature and precipitation (Fig. 4), it is evident that during the days and months of sampling, precipitation was little or none (monthly average from 1.3 mm in January to 7.4 mm in April) whereas monthly average temperature tended to increase from 21.5 °C to 23.3 °C from January to March and from 24.0 to 24.5 °C from April to June. These daily temperature records showed significant monthly differences (ANOVA, F = 67.1, df5 = 5, df2 = 79.8, p < 0.001). There were no significant monthly fluctuations with respect to precipitation (ANOVA, F = 1.78, df5 = 5, df2 = 78.8, p > 0.05). The species richness tended to be higher in April and June in all 3 vegetation covers (Fig. 5) whereas abundance was higher in the forest in March and higher for pastures and crops during April and June. Species diversity (Shannon and Simpson) in the forest was higher in March whereas in both crops and pastures it was higher in June (Fig. 5).
Based on the Jaccard similarity index, crop and pastures presented a greater degree of similarity (Fig. 6), that is, a greater number of shared species. The forest presented the greatest dissimilarity in species composition with respect to the crop and the pasture, having a greater number of unique species (D. truncatus, D. ebraccatus, L. poecilochilus, and L. savagei), which are shown in Table 1.
It was observed that the microhabitat most used by anurans in the forest was leaf litter. The species most associated with this type of microhabitat were D. truncatus and L. poecilochilus (Fig. 7); these species were only recorded in this coverage (Table 1). In the pasture, the highest record of species was found in bare soils and jagüeyes, with L. fuscus, L. insularum, and B. pugnax being the most associated with the latter, while R. horribilis and R. humboldti were observed mainly in bare soil (Fig. 8). In addition, these species presented the highest number of records of individuals in this coverage (Table 1). Finally, in the crop coverage, the microhabitat with the highest number of anuran records was bare soil (Fig. 9), this microhabitat was used most frequently by R. humboldti, R. horribilis, and P. brachyops which were the species with the highest number of individuals recorded; this microhabitat was also used by E. panamensis, which was the only species present in this cover (Table 1).
Discussion
In this study, 19 species of anurans and 1 casual record were identified, for a total of 20 species, this being a slightly smaller number than the 21 species recorded in the SFF Los Colorados 2018-2023 Management Plan (Jiménez et al., 2018). Craugastor raniformis (Boulenger, 1896), Pseudopaludicola pusilla (Ruthven, 1916), and Lectodactylus fragilis (Brocchi, 1877) were not observed in our study, possibly due to lack of sampling in some areas of the SFF Los Colorados. Their occurrence cannot be ruled out, since they were recorded by Acosta-Galvis (2012) in the Montes de María. This study reports C. raniformis in the forest, in ravines (on rocks), on leaf litter, and in shrubby vegetation; P. pusilla in crop areas and on the edge of plain forests, on sandy substrates and in cracks after rains; L. fragilis in flat areas, around seasonal ponds, and near swamps.
Scarthyla, D. ebraccatus, and L. savagei are added to the anuran fauna of the SFF Los Colorados, which shows that it is necessary to continue carrying out studies in the subregions of STDF, including the protected areas of the plains of the Caribbean region, valleys of the Magdalena and Cauca Rivers, Catatumbo, and enclaves of the Patía Valley. Amphibian diversity is poorly known due to the lack of biological studies (Urbina-Cardona et al., 2014).
Sampling carried out in the first 3 months of the year (January, February, and March) regularly corresponds to the dry season (Rangel & Carvajal-Cogollo, 2012). However, rains occurring in these months is a consequence of the effects of the La Niña phenomenon in Colombia for 2022 (Guzmán-Ferraro & García, 2022).
Figure 4. Histogram and box plot of daily ambient temperature across the sampling months; temperature (A, B), rainfall (C).
Increases were observed in the specific richness and in the recorded number of individuals as rainfall increased (especially in April and June), so it was considered that the rainfall regime prior to sampling played an important role in the observation of anurans. These increases in richness and mainly in the number of individuals are attributed to higher activity and the reproductive strategies of some species, which take advantage of the rains to reproduce and lay eggs in temporary ponds. The rains caused greater activity and detectability of some species that were observed vocalizing in small ponds that had formed and cow dams.
Figure 5. Monthly trend of average richness (A), abundance (B), and diversity (C, D) for each vegetation cover type.
Only some amplexuses were recorded but we did not record nesting or reproduction events. Some of the species have explosive activity, which generates an increase in the number of individuals, as is the case of R. horribilis and other species (Vargas-Salinas et al., 2019); some other species vocalizing included Engytomops pustulosus in some ponds and Dendropsophus microcephalus in the emerging vegetation around the cow dams. However, it is worth mentioning that the frequency and intensity of the La Niña phenomenon due to climate change could alter the reproductive times of anurans, causing many species to have early reproduction, which would bring about temporal overlaps of the species that would generate changes in the structure of the assembly (Lawyer & Morin, 1993).
Figure 6. Jaccard similarity dendrogram for the anuran samples from the SFF Los Colorados.
As we expect, there were differences of species richness among vegetation covers. The forest recorded greater richness of anurans than the productive systems (pasture and crops). This is mainly attributed to the greater availability of humid microhabitats and the vegetaion complexity, since there are species that require dense vegetation cover and abundant leaf litter. For example, the oviposition of D. truncatus occurs in humid leaf litter (Cárdenas-Ortega et al., 2019), so different studies record it abundantly in forested areas (Burbano-Yandi et al., 2016; De la Ossa et al., 2016, 2011; Posso-Peláez et al., 2017). On the other hand, pasture and crop are covers with less complexity in the vegetal structure, generating changes in the composition of the anuran assemblages (Cortés-Gómez et al., 2013), such as the reduction in richness, which is closely linked to the reproductive modes of each species (Almeida-Gomes & Rocha, 2015). These same changes in richness in covers with different degrees of disturbance have been recorded in different studies carried out in the Middle Magdalena Valley, in Meta, and in Florencia (Burbano-Yandi et al., 2016; Cáceres-Andrade & Urbina-Cardona, 2009; Duarte-Marín et al., 2018). However, total and monthly average species richness and diversity tended to be higher in the pastures and the lowest in the forests, which registered the greatest monthly variation, recorded the higher species richness and diversity in April and June. The greater diversity of species recorded in pastures may be associated with the lower complexity of the vegetation structure of this habitat, which allows anurans to be easier to detect, while the greater structural complexity of forests and crops decreases detectability of the anurans. Additionally, the presence of jagüeyes in pastures, as sites with availability of water and constant humidity necessary for the survival of the anurans, contributed to the greatest number of records of individuals in this vegetation coverage. Leptodactylus fuscus, L. insularum, and B. pugnax had greater abundance in cow dams of the pasture, since they make postures close to bodies of water (Carvajal-Cogollo et al., 2019; Ortega-Chinquilla et al., 2019).
Figure 7. Use of microhabitats by species in the forest cover of the SFF Los Colorados.
Jagüeyes are considered important for many species in disturbed areas because they permanently provide water resources, which can be used to increase water uptake and reduce evaporation rates (Urbina-Cardona et al., 2014), in addition to be used by species with reproductive modes associated with this resource (Cardozo & Caraballo, 2017). On the other hand, bare soils were mostly used by R. horribilis and R. humboldti, which are species that are commonly found in disturbed areas (Acosta-Galvis, 2012), these have physical characteristics (tuberculated skin) and physiological characteristics that allow them to adapt to exploit this microhabitat (Cáceres-Andrade & Urbina-Cardona, 2009), for this reason, they were found with greater abundance in pastures and crops.
Dendrobates truncatus, D. ebraccatus, L. poecilochilus, and L. savagei are species that were recorded only in the forest, similar to what has been reported by other anuran assemblage studies, where they are not only recorded in forested areas, but also in wetlands (Acosta-Galvis et al., 2006; Angarita et al., 2015; Burbano-Yandi et al., 2016). On the other hand, S. vigilans was only recorded in pastures; in this study the species was observed mainly around bodies of water, specifically on emergent vegetation, in sympatry with D. microcephalus (Muñoz-Guerrero et al., 2007; Fonseca- Pérez et al., 2017). Finally, E. panamensis was only present in crops, although in the study carried out by Blanco-Torres et al. (2015), it was also recorded in pastures. This is a species identified as a leaf miner (Cuentas et al., 2002), which has possibly been the reason why it was observed near the cracks produced by cassava plantations. It is due to all the above that the species similarity analysis showed that the forest differs with respect to the other 2 covers, which are noticeably more similar to each other. Just as we expected, the forest differs in species composition.
Figure 8. Use of microhabitats by species in the pasture cover of the SFF Los Colorados.
The species diversity index values reported in our study are similar to those obtained in a nearby area located in Meta, Colombia (Cáceres-Andrade and Urbina-Cardona, 2009) where they reported values of 1.4 for humid forest, 1.43 in pastures and 1.9 in sugarcane crops. On the other hand, Román-Palacios et al. (2016) estimated a low Shannon-Wiener index in the Magdalena Medio for forest and quarry (0.92 and 1.74, respectively), while for the lake, they estimated an intermediate diversity (2.03). The forest value was very far from that estimated in this work, which may indicate that the SFF Los Colorados forest has an important conservation status that benefits the anurafauna.
On the other hand, Simpson’s index estimated high dominance for the 3 plant covers. This dominance may be associated with the microhabitats they offer; for example, the forest offered important microhabitats (numerous ponds and abundant leaf litter) for the development of E. pustulosus and D. truncatus, which made these species dominant in this cover. On the other hand, the crop was dominated mainly by P. brachyops, R. humbolbti, and R. horribilis; these species have terrestrial habits, tolerant to landscape transformations and abundant open environments (Acosta-Galvis, 2012; Rodríguez-Molina, 2004). Finally, the pasture was dominated by species of the genus Rhinella and Leptodactylus, where the latter has reproductive modes associated with foam nests, allowing them to conquer and be abundant in anthropized environments (Alcaide et al., 2012).
Jaccard’s similarity analysis for the anurans of the SFF Los Colorados indicates a grouping between pasture and crop cover due to the percentage of shared species (66.6%). This result may be associated with the fact that both covers are intervened areas, with a vegetation structure different from that of the forest and host generalist species (e.g., P. brachyops, B. pugnax, L. fuscus) that can share in greater quantities, while the forest, due to the resources it offers, may have species that do not tolerate landscape transformations (e.g., D. truncatus), being restricted only to forested areas (Cáceres-Andrade & Urbina-Cardona, 2009).
Figure 9. Use of microhabitats by species in the crop cover of the SFF Los Colorados.
The results of this research indicate that the transformation of the landscape because of the agricultural economy of the Montes de María, based mainly on cultivation and the raising of animals (Aguilera-Díaz, 2013), generated changes in the wealth, abundance, composition and use of microhabitats in anuran assemblages. Therefore, this knowledge is important to create concrete tools for the management and conservation of these organisms in the protected area and its surroundings, such as maintaining native vegetation and layers of leaf litter in productive systems, conserving lentic and lotic water sources, and reducing the use of agrochemicals, among others (Urbina-Cardona et al., 2015).
This research constitutes the baseline to evaluate the long-term response of anurans to ecological restoration processes and initiatives led by the SFF Los Colorados team in transformed areas of the protected area. This research constitutes the baseline to evaluate the long-term response of anurans to ecological restoration processes and initiatives led by the SFF Los Colorados team in transformed areas of the protected area. Results that could be useful in future studies where reference ecosystems (conserved areas) and disturbed areas in the process of restoration are used, to determine if these protected areas are achieving the expected objectives and if they are contributing to the conservation of anurans (Urbina-Cardona et al., 2015). Additionally, this study updates the list of anuran species in the protected area, pointing out those to a specific coverage and those shared among covers (forest, pasture, and crops), which can be useful to define those that may be vulnerable to fragmentation of the habitat or to be included as conservation target values (VOC) of the SFF Los Colorados in the construction of future Management Plans.
Acknowledgments
We thank the Hydrobiology research group of the University of Cartagena for providing their equipment for the development of this research. Likewise, to the University of Cartagena, for the financial support through resolution number 01878 of 2022. To IDEAM for providing information on the environmental variables for 2022. To Gabriel R. Navas-S, Dr. Andrés García, and Vivarium del Caribe for financial support and suggestions for carrying out this study. Likewise, to the technical and administrative team of the Los Colorados Flora and Fauna Sanctuary for their support in sampling, logistics, loan of facilities, and management of the research guarantee. To Joselin Castro-Palacios for his support in the implementation of the methodology of this work and to Adolfo A. Mulet-Paso for his suggestions in the identification of the species. To David Gernandt, for his revision to the text which improved substantially.
References
Acosta-Galvis, A. R. (2012). Anfibios de los enclaves secos del área de influencia de los montes de María y la ciénaga de La Caimanera, departamento de Sucre, Colombia. Biota Colombiana, 13, 211–255.
Acosta-Galvis, A. R. (2021). Lista de anfibios de Colombia. BATRACHIA. Retrieved on June 17, 2021 from: http://www.batrachia.com
Acosta-Galvis, A. R., Huertas-Salgado, C., & Rada, M. (2006). Aproximación al conocimiento de los anfibios en una localidad del Magdalena medio (Departamento de Caldas, Colombia). Revista de la Academia Colombiana de Ciencias Exactas, Físicas y Naturales, 30, 291–303. https://doi.org/10.18257/raccefyn.30(115).2006.2250
Acuña-Vargas, J. C. (2016). Anfibios y reptiles asociados a cinco coberturas de la tierra, municipio de Dibulla, La Guajira, Colombia. Acta Zoológica Mexicana, 32,133–146. https://doi.org/10.21829/azm.2016.322940
Aguilera-Díaz, M. M. (2013). Montes de María: una subregión de economía campesina y empresarial. Cartagena: CEER.
Aguirre-León, G. (2011). Métodos de estimación, captura y contención de anfibios y reptiles. Manual de Técnicas para el estudio de la Fauna. In S. López-González y C. López-González (Eds.), Manual de técnicas para el estudio de fauna (pp. 61–85). Querétaro: Universidad Autónoma de Querétaro, A.C.
Alcaide, A. P., Ponssa, M. L., Alcaide, F. P., & Alcaide, M. (2012). Histología de ovario en hembras vitelogénicas de Leptodactylus latinasus (Anura, Leptodactylidae). Acta Zoológica Lilloana, 56, 44–53.
Almeida-Gomes, M., & Rocha, C. F. (2015). Habitat loss reduces the diversity of frog reproductive modes in an Atlantic Forest fragmented landscape. Biotropica, 47, 113–118. https://doi.org/10.1111/btp.12168
Andrade, M. G. (2011). Estado del conocimiento de la biodiversidad en Colombia y sus amenazas. Consideraciones para fortalecer la interacción ciencia-política. Revista de la Academia Colombiana de Ciencias Exactas, Físicas y Naturales, 35, 491–508.
Angarita, M. O., Montes-Correa, A. C., & Renjifo, J. M. (2015). Amphibians and reptiles of an agroforestry system in the Colombian Caribbean. Amphibian and Reptile Conservation, 8, 33–52.
Blanco, A., & Bonilla, M. (2010). Partición de microhábitats entre especies de Bufonidae y Leiuperidae (Amphibia: Anura) en áreas con bosque seco tropical de la región Caribe-Colombia. Acta Biológica Colombiana, 15,47–60.
Blanco-Torres, A., Duré, M., & Bonilla, M. A. (2015). Observaciones sobre la dieta de Elachistocleis pearsei y Elachistocleis panamensis en dos áreas intervenidas de tierras bajas del norte de Colombia. Revista Mexicana de Biodiversidad, 86,538–540. https://doi.org/10.1016/j.rmb.2015.04.031
Ballesteros-Correa, J., Vidal-Pastrana, C., & Ortega-León, A. M. (2019). Anfibios de Córdoba, Colombia. Córdoba: Fondo Editorial de Córdoba.
Burbano-Yandi, C. E., Gómez-Díaz, M. A., Gómez-Figueroa, A., Velásquez-Trujillo, D. A., & Bolívar-García, W. (2016). Ensamblaje de anfibios presentes en un bosque seco y en sistemas productivos, Valle Medio del Magdalena, Victoria y La Dorada, Caldas, Colombia. Revista de Ciencias, 20, 81–93.
Cáceres-Andrade, S. P., & Urbina-Cardona, J. N. (2009). Ensamblajes de Anuros de sistemas productivos y bosques en el Piedemonte Llanero, departamento del Meta, Colombia. Caldasia, 31, 175–195.
Cárdenas-Ortega, M. S., Gutiérrez-Cárdenas, P. D., Cifuentes-Ortegón, M., & Patiño-Gallego, A. M. (2019). Dendrobates truncatus (Cope, 1861) Rana venenosa de rayas amarillas. Catálogo de Anfibios y Reptiles de Colombia, 5, 32–41.
Cardozo, J., & Caraballo, P. (2017). Fauna anura (Amphibia: Anura) asociada a jagüeyes en dos localidades de la región Caribe colombiana. Revista Colombiana de Ciencia Animal-RECIA, 9, 39–47. https://doi.org/10.24188/recia.v9.nS.2017.519
Carvajal-Cogollo, J. E., Bernal-González, V., Paternina-Hernández, A., Muñoz-Ávila, J. N., & Vargas-Salinas, F. (2019). Uso de hábitat y reglas de ensamble: patrones y mecanismos. In M. H. Restrepo-Domínguez, E. Vera-López, Y. Bolívar-Suárez, S. G. Numpaque-Piracoca, O. Y. Acuña-Rodríguez, Z. Z. Ojeda-Pérez et al. (Eds.), Biología de los anfibios y reptiles en bosque tropical del norte de Colombia (pp. 297–338). Tunja: Editorial UPTC.
Clarke, K. R., Gorley, N. R., Somerfield, P. J., & Warwick, R. M. (2014). Change in marine communities: an approach to statistical analysis and interpretation. Plymouth: PRIMER-E.
Colwell, R. K., & Coddington, J. A. (1994). Estimating terrestrial biodiversity through extrapolation. Philosophical Transactions of the Royal Society of London. Series B: Biological Sciences, 345,101–118. https://doi.org/10.1098/rstb.1994.0091
Conell, J. H. (1978). Diversity in tropical rain forests and coral reefs. Science, 199,1302–1310. http://dx.doi.org/10.1126/science.199.4335.1302
Cortés-Gómez, A. M., Castro-Herrera, F., & Urbina-Cardona, J. N. (2013). Small changes in vegetation structure create great changes in amphibian ensembles in the Colombian Pacific rainforest. Tropical Conservation Science, 6, 749–769. https://doi.org/10.1177/194008291300600604
Cortez, F. C., Suárez-Mayorga, A. M., & López-López, F. J. (2006). Preparación y preservación de material científico. In A. Angulo, J. V. Rueda-Almonacid, J. V. Rodríguez-Mahecha y E. La Marca (Eds.), Técnicas de inventario y monitoreo para los anfibios de la región tropical andina (pp. 173–218). Bogotá: Panamericana Formas e Impreso S.A.
Cristal, A., Sánchez, E., Romero, J., Leyva, J., Andrade, W., & Buelvas, C. (2020). ¿Qué hemos logrado con el proyecto de Conectividades Socio-Ecosistémicas? La evolución de la metodología de las 4Ps: avances y retos. In A. Cristal, M. Peña y J. Ferrer-Sotelo (Eds.), El proyecto de conectividades Socio-Ecosistémicas en los Montes de María, 2013–2020 (pp. 63–87). Bogotá-Colombia: Fundación Herencia Ambiental Caribe.
Crump, M., & Scott, N. (2001). Relevamiento por encuentros visuales. In W. R. Heyer, M. A. Donnelly, R. W. Diarmid, L. C. Hayek, & M. S. Foster (Eds.), Medición y monitoreo de la diversidad biológica: métodos estandarizados para anfibios (pp. 80–86). Chubut, Argentina: Editorial Universitaria de la Patagonia.
Cuentas, D., Borja, R., Lynch, J. D., & Renjifo, J. M. (2002). Anuros del departamento del Atlántico y norte de Bolívar. Barranquilla: Editorial Cencys.
De Sá, R. O. (2005). Crisis global de biodiversidad: importancia de la diversidad genética y la extinción de anfibios. Agrociencia, 9, 513.
De la Ossa, J., Contreras-Gutiérrez, J. C., & Campillo-Castro, J. (2011). Censo de Dendrobates truncatus (Anura, Dendrobatidae) en la reserva forestal protectora Serranía de Coraza, Montes de María, Sucre, Colombia. Revista Colombiana de Ciencia Animal-RECIA, 3, 339–343. https://doi.org/10.24188/recia.v3.n2.2011.407
Díaz, S., Fargione, J., Chapin, F. S., & Tilman, D. (2006). Biodiversity loss threatens human well-being. Plos Biology, 4, 1300–1305. https://doi.org/10.1371/journal.pbio.0040277
Duarte-Marín, S., González-Acosta, C., & Vargas-Salinas, F. (2018). Estructura y composición de ensamblajes de anfibios en tres tipos de hábitat en el Parque Nacional Natural Selva de Florencia, Cordillera Central de Colombia. Revista de la Academia Colombiana de Ciencias Exactas, Físicas y Naturales, 42, 227–236. https://doi.org/10.18257/raccefyn.631
Dunn, E. R. (1944). Los géneros de anfibios y reptiles de Colombia. Caldasia, 2,497–529.
Echeverry, M. A., & Rodríguez, J. M. (2006). Análisis de un paisaje fragmentado como herramienta para la conservación de la biodiversidad en áreas de bosque seco y subhúmedo tropical en el municipio de Pereira, Risaralda Colombia. Scientia et Technica, 12, 405–410.
Etter, A., Andrade, A., Saavedra, K., Amaya, P., Cortés, J., Pacheco, C. et al. (2017). Lista Roja de Ecosistemas de Colombia, 2,1–6. https://doi.org/10.13140/RG.2.2.10861.08165
Etter, A., McAlpine, C., & Possingham, H. (2008). Historical patterns and drivers of landscape change in Colombia since 1500: a regionalized spatial approach. Annals of the Association of American Geographers, 98, 2–23. https://doi.org/10.1080/00045600701733911
Fonseca-Pérez, K. A., Molina, C., & Tárano, Z. (2017). Diet of Dendropsophus microcephalus and Scarthyla vigilans (Anura: Hylidae) at a locality in north-western Venezuela with notes on microhabitat occupation. Papéis Avulsos de Zoologia, 57,93–104. https://doi.org/10.11606/0031-1049.2017.57.07
Fox, B. J., Taylor, J. E., Fox, M. D., & Williams, C. (1997). Vegetation changes across edges of rainforest remnants. Biological Conservation, 82, 1–13. https://doi.org/10.1016/S0006-3207(97)00011-6
Galván-Guevara, S., Ballut-Dajud, G., & De La Ossa, J. (2015). Determinación de la fragmentación del bosque seco del arroyo Pechelín, Montes de María, Caribe, Colombia. Biota Colombiana, 16, 149–157.
García, R. J. C., Castro, H. F., & Cárdenas, H. H. (2005). Relación entre la distribución de anuros y variables del hábitat en el sector La Romelia del Parque Nacional Natural Munchique (Cauca, Colombia). Caldasia, 27, 299–310.
García, H., Corzo, G., Isaacs, P., & Etter, A. (2014). Distribución y estado actual de los bosques remanentes del bioma de bosque seco tropical en Colombia: insumos para su gestión. In C. Pizano, & H. García (Eds.), El bosque seco tropical en Colombia (pp. 228–251). Bogotá: Instituto de Investigación de Recursos Biológicos Alexander von Humboldt.
Guzmán-Ferraro, M., & García, G. (2022). El fenómeno de La Niña persistirá hasta casi la mitad del año. CVC. Retrieved on January 14th, 2022 from: http://www.cvc.gov.co/boletin-prensa-007-2022
Hammer, Ø., Harper, D. A., & Ryan, P. D. (2001). PAST: Paleontological statistics software package for education and data analysis. Palaeontologia Electronica, 4, 1–9.
Herazo, F., Mercado, J., & Mendoza, H. (2017). Estructura y composición florística del bosque seco tropical en los Montes de María (Sucre-Colombia). Ciencia en Desarrollo, 8,71–8.
Instituto Alexander von Humboldt. (1998). El bosque seco tropical en Colombia B-sT. Grupo de Exploraciones y Monitoreo Ambiental. Bogotá, Colombia.
IDEAM (Instituto de Hidrología, Meteorología y Estudios Ambientales), IGAC (Instituto Geográfico Agustín Codazzi), & CORMAGDALENA (Corporación Autónoma Regional del río Grande de La Magdalena). (2008). Mapa de cobertura de la Tierra Cuenca Magdalena-Cauca: Metodología CORINE Land Cover adaptada para Colombia a escala 1:100.000. Bogotá: IDEAM/ IGAC/ CORMAGDALENA.
Jiménez, B., De la Rosa, N., & Naranjo, D. (2018). Plan de manejo del Santuario de Flora y Fauna Los Colorados. Parques Nacionales Naturales de Colombia. Santa Marta: Dirección Territorial Caribe.
Laurance, W. F., & Gascon, C. (1997). How to creatively fragment a landscape. Conservation Biology, 11, 577–579.
Lawler, S., & Morin, P. (1993). Temporal overlap, competition, and priority effects in larval anurans. Ecology, 71, 174–182. https://doi.org/10.2307/1939512
Magurran, A. (2004). Measuring biological diversity. Malden: Blackwell Publishing.
Manzanilla, J., & Péfaur, J. E. (2000). Consideraciones sobre métodos y técnicas de campo para el estudio de anfibios y reptiles. Revista de Ecología Latinoamericana, 7,17–30.
Marín, A. M., Ríos, L., Ríos, L., & Almario, J. (2017). Impacto de la actividad ganadera sobre el suelo en Colombia. Ingeniería y Región, 17, 1–12. https://doi.org/10.25054/issn.2216-1325
Meza-Tílvez, K., Mulet-Paso, A., & Zambrano-Cantillo, R. (2018). Fauna del Jardín Botánico de “Guillermo Piñeres” de Cartagena, Turbaco, Colombia: Anfibios y reptiles. Versión 1. Chicago, Illinois: Field Museum.
Moreno, C. E. (2001). Métodos para medir la biodiversidad. Volumen 1. Zaragoza: CYTED, ORCYT/ UNESCO & SEA.
Muñoz-Guerrero, J., Serrano, V. H., & Ramírez-Pinilla, M. P. (2007). Uso de microhábitat, dieta y tiempo de actividad en cuatro especies simpátricas de ranas hílidas neotropicales (Anura: Hylidae). Caldasia, 29, 413–425.
McDiarmid, R. W. (1994). Preparing amphibians as scientific specimens. In W. R. Heyer, M. A. Donnelly, R. W. McDiarmid, L. C. Hayek, & M. S. Foster (Eds.), Measuring and monitoring biological diversity. standard methods for amphibians (pp: 289–296). Washington D.C.: Smithsonian Institution Press.
O’Malley, B. (2007). Anatomía y fisiología clínica de animales exóticos. Zaragoza: Servet.
Ortega-Chinquilla, J., Méndez-Narváez, J., Carvajalino-Fernández, J., & Galindo-Uribe, D. (2019). Ecofisiología. In F. Vargas-Salinas, J.A. Muñoz-Avila & M.E. Morales-Puente (Eds.), Biología de los anfibios y reptiles en bosque tropical del norte de Colombia (pp. 297–338). Tunja: Editorial UPTC.
Pizano, C., Gonzáles, R., Gonzáles, M., Castro-Lima, R., Rodríguez, N., Idárraga, A. et al. (2014). Plantas de los bosques secos de Colombia. In C. Pizano, & H. García (Eds.), El bosque seco tropical en Colombia (pp. 228–251). Bogotá: Instituto de Investigación de Recursos Biológicos Alexander von Humboldt.
Posso-Peláez, C., Blanco-Torres, A., & Gutiérrez-Moreno, L. C. (2017). Uso de microhábitats, actividad diaria y dieta de Dendrobates truncatus (Cope, 1861) (Anura: Dendrobatidae) en bosque seco tropical del norte de Colombia. Acta Zoológica Mexicana, 33, 490–502.
Rangel, C. J. O., & Carvajal-Cogollo, J. E. (2012). Clima de la región Caribe Colombiana. In J. O. Rangel (Ed.), Colombia diversidad biótica XII: la región Caribe de Colombia (pp. 67–129). Bogotá: Instituto de Ciencias Naturales.
Re:wild, Synchronicity Earth, IUCN SSC Amphibian Specialist Group. (2023). State of the World’s Amphibians: The Second Global Amphibian Assessment. Texas: Re:wild. Electronic source at the IUCN website
Rodríguez-Molina, C. R. (2004). Reproducción de Pleurodema brachyops (Anura: Lectodactylidae) en los llanos del Estado Apure, Venezuela. Memoria de la Fundación La Salle Ciencias Naturales, 2002, 117–125.
Rodríguez, G. M., Banda, K., Reyes, S. P., & Estupiñán-González, A. C. (2012). Lista comentada de las plantas vasculares de bosques secos prioritarios para la conservación en los departamentos de Atlántico y Bolívar (Caribe colombiano). Biota Colombiana, 13, 7–39.
Rojas, R. R., & Pérez-Peña, P. E. (2018). Evidencia preliminar del efecto borde en anfibios de la Reserva Nacional Pucacuro, al norte de la Amazonía peruana. Revista del Instituto de Investigaciones de la Amazonía Peruana, 27, 55–67.
Romero, H. G. (2013). Deforestación en Colombia: retos y perspectivas. In F. Dane (Ed.), El desafío del desarrollo sustentable en América Latina (pp. 123–142). Río de Janeiro: SOPLA.
Román-Palacios, C., Fernández-Garzón, S., Hernández, M., Ishida-Castañeda, J., Gallo-Franco, W., & Bolívar-García, A. (2016). Uso de microhábitat por anuros en un fragmento de bosque seco intervenido del Magdalena Medio, Guarinocito, Caldas. Boletín Científico Centro de Museos Museo de Historia Natural, 20, 181–196. https://doi.org/10.17151/bccm.2016.20.2.14
Salvador, J., & Gómez, D. (2018). Reptiles y anfibios del departamento del Atlántico, Colombia. Versión 3. Bogotá, Colombia: Field Museum-Museo de Historia Natural ANDES.
Stuart, S. N., Chanson, J., Cox, N.A., & Young, B. E. (2006). Estado global de los anfibios. In A. Angulo, J. V. Rueda-Almonacid, J. V. Rodríguez-Mahecha, & E. La Marca (Eds.), Técnicas de inventario y monitoreo para los anfibios de la región tropical andina (pp. 19–41). Bogotá: Panamericana Formas e Impreso S.A.
Simmons, J. E., & Muñoz-Saba, Y. (Eds.). (2005). Cuidado, manejo y conservación de las colecciones biológicas. Bogotá: Universidad Nacional de Colombia.
Thompson, P. L., Rayfield, B., & González, A. (2017). Loss of habitat and connectivity erodes species diversity, ecosystem functioning, and stability in metacommunity networks. Ecography, 40, 98–108. https://doi.org/10.1111/ecog.02558
Urbina-Cardona, J. N., Olivares-Pérez, M., & Reynoso, V. H. (2006). Herpetofauna diversity and microenvironment correlates across a pasture-edge-interior ecotone in tropical rainforest fragments in the Los Tuxtlas Biosphere Reserve of Veracruz, Mexico. Biological Conservation, 132,61–75. https://doi.org/10.1016/j.biocon.2006.03.014
Urbina-Cardona, J. N., Arturo-Navas, C., Gonzales, I., Gómez-Martínez., M. J., Llano-Mejía, J., Medina-Rangel, G. F. et al. (2014). Determinantes de la distribución de los anfibios en el bosque seco tropical de Colombia: herramientas para su conservación. In H. Pizano, & H. García (Eds.), El bosque seco tropical en Colombia (pp. 169–195). Bogotá: Instituto de Investigación de Recursos Biológicos Alexander von Humboldt.
Urbina-Cardona, J. N., Bernal, E. A., Giraldo-Echeverry, N., & Echeverry-Alcendra, A. (2015). El monitoreo de herpetofauna en los procesos de restauración ecológica: indicadores y métodos. Monitoreo a procesos de restauración ecológica, aplicado a ecosistemas terrestres. Bogotá: Instituto de Investigación de Recursos Biológicos Alexander von Humbold.
Vargas-Salinas, F., Angarita-Sierra, T., Ospinal, L. A., Rocha-Úsuga, A., & Rueda-Solano, L. (2019). Comunicación y ecología reproductiva. In F. Vargas-Salinas, J.A. Muñoz-Avila & M.E. Morales-Puente (Eds.), Biología de los anfibios y reptiles en bosque tropical del norte de Colombia (pp. 297–338). Tunja: Editorial UPTC.
Vargas, S. F., & Bolaños, L. M. E. (1999). Anfibios y reptiles presentes en hábitats perturbados de selva lluviosa tropical en el bajo Anchicayá, Pacífico colombiano. Revista de la Academia Colombiana de Ciencias Exactas, Físicas y Naturales, 23,499–511.
Villareal, H. M., Álvarez, M., Córdoba-Córdoba, S., Escobar, F., Fagua, G., Gast, F. et al. (2004). Manual de métodos para el desarrollo de inventarios de biodiversidad. Bogotá: Instituto de Investigación de Recursos Biológicos Alexander von Humboldt.
Zug, G. R., Vitt, L., & Caldwell, J. P. (2001). Herpetology: an introductory biology of amphibians and reptiles. San Diego: Academic Press.
Association of Myianoetus sp. (Acari: Histiostomatidae) with necrophagous fly Compsomyiops fulvicrura (Diptera: Calliphoridae), in the Prepuna ecoregion (Jujuy: Argentina)
María Laura Fernández-Salinas a, *, Marcia Luciana Matoz-Fernández b
a Universidad Nacional de Jujuy, Instituto de Biología de la Altura, Avenida Bolivia 1661, 4600 San Salvador de Jujuy, Jujuy, Argentina
b Universidad Nacional de Mar del Plata, Laboratorio de Zoonosis Parasitarias, Funes 3350, 7600 Mar del Plata, Buenos Aires, Argentina
*Corresponding author: mfernandez@inbial.unju.edu.ar (M.L. Fernández-Salinas)
Received: 31 January 2024; accepted: 10 June 2024
Abstract
The genus Myianoetus Oudemans(Acari: Histiostomatidae) is commonly associated with carrion, utilizing flies (Diptera) from various families as a means of dispersal through phoresy. The objective of this paper is to present a new association between Myianoetus sp. mites and Calliphoridae flies and discuss its relevance in forensic sciences. Samples were collected in 3 locations in the Prepuna ecoregion of Jujuy, Argentina. Specimens were captured using necrotraps baited with cow lung. Flies carrying phoretic mites were separated and identified to a specific level, while mites were counted and identified at the lowest possible taxonomic level. Compsomyiops fulvicrura (Robineau-Desvoidy) (Diptera: Calliphoridae) was the only species that presented attached mites, with an average intensity of 12.26 mites per fly. The mites carried by C. fulvicrura were identified as deutonymphs of Myianoetus sp., with a prevalence of 2.56% of infested flies. Significant differences in the abundance of flies with mites were observed between locations and seasons. This article represents the first contribution to knowledge on the specific association between Myianoetus sp. and C. fulvicrura. These findings in forensic ecology are relevant for their potential contribution and application in the development of more precise methods in specific forensic cases.
Keywords: Astigmata; Diptera; Forensic Acarology; Phoresy; New report
© 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/).
Asociación de Myianoetus sp. (Acari: Histiostomatidae) con la mosca necrófaga Compsomyiops fulvicrura (Diptera: Calliphoridae), en la ecoregión Prepuna (Jujuy: Argentina)
Resumen
El género Myianoetus Oudemans (Acari: Histiostomatidae) suele asociarse a la carroña utilizando moscas (Diptera) de distintas familias como medio de dispersión, a través de la foresia. El objetivo de este trabajo fue presentar una nueva asociación entre Myianoetus sp. con moscas Calliphoridae y discutir su alcance dentro de las ciencias forenses. Las muestras se recolectaron en 3 localidades de la Prepuna jujeña, Jujuy, Argentina. Los especímenes se capturaron mediante necrotrampas cebadas con pulmón vacuno. Las moscas con ácaros se separaron y determinaron a nivel específico; los ácaros fueron numerados e identificados al nivel taxonómico más bajo posible. Compsomyiops fulvicrura (Diptera: Calliphoridae) fue la única especie que presentó ácaros adheridos, con una intensidad media de 12.26 ácaros por mosca. Los ácaros fueron identificados como deutoninfas de Myianoetus sp. y se determinó una prevalencia de 2.56% de moscas infestadas. Se observaron diferencias significativas en la abundancia de moscas con ácaros entre las localidades y estaciones analizadas. Este artículo representa el primer aporte al conocimiento sobre la asociación específica entre Myianoetus sp. y C. fulvicrura. Estos hallazgos sobre ecología forense son relevantes por su potencial contribución y aplicación al desarrollo de métodos más precisos en casos forenses determinados.
Palabras clave:Astigmata; Diptera; Acarología forense; Foresia; Nuevo reporte
Introduction
Carcasses present limited and ephemeral biocenosis made up of diverse organisms that often comprise complex food webs (Braig & Perotti, 2009; Perotti et al., 2010). Many Diptera species actively participate in the cadaveric decomposition process in which the Calliphoridae and Sarcophagidae families, along with Coleoptera are often investigated because of their large number, persistence and capacity to act as hosts to diverse mites that use them for dispersion by phoresis (Camerick, 2010; Perotti & Braig, 2009; Perotti et al., 2010).
Mites present morphological and physiological adaptations to serve phoresy during adult and nymphal stages. These adaptations are documented in the order Mesostigmata, in the suborder Prostigmata and in the infraorder Astigmatina (Oribatida) (Perotti et al., 2010). Astigmata mites are specialists in irregular or ephemeral habitats which they colonize through a deutonymphal heteromorphic stage known as hypopus which is specialized for phoresy (OConnor, 2009). Astigmatid deutonymphs are morphologically simplified, have lost the mouth and chelicerae, have greatly reduced the remainder of the gnathosoma, and have suckers on the paraproctal region for efficient phoretic attachment. The body is strongly dorsoventrally flattened, heavily sclerotized and much more resistant to desiccation than other stages of the life cycle (Farish & Axtell, 1971; OConnor, 1982). The conditions needed to reach this stage may involve genetic factors and physicochemical factors from the environment (Greenberg & Carpenter, 1960).
Astigmatid mites are particularly important for the 3 areas of forensic entomology: urban, stored product pests and medico-legal (Catts & Goff, 1992; Perotti & Braig, 2019). Nevertheless, they often go unnoticed because of their small size. Moreover, their analysis is limited because of difficulties in species identification, lack of specific knowledge and misuse of forensic methodology (OConnor, 2009; Perotti et al., 2010). Numerous species of mites are compulsory or facultative inhabitants of carrion. They are found not only in legal cases that involve human carcasses (Pimsler et al., 2016; Rai et al., 2020; Russell et al., 2004; Saloña-Bordas & Perotti, 2015); but also, in experimental studies concerning faunal succession in animal remains (Arnaldos et al., 2005; Barton et al., 2014; Centeno & Perotti, 1999; Heo et al., 2021).
In Argentina, the only record of the presence of phoretic mites associated with decomposing remains were the preliminary observations of Centeno and Perotti (1999), in which they found mites of the genus Myianoetus Oudemans (Astigmata: Histiostomatidae) associated with a specimen of Morellia sp. (Muscidae). In order to contribute to the further study on phoretic relations between mites and arthropods, this paper presents a new association between mites and Diptera from the Calliphoridae family in Prepuna of Jujuy, Argentina, and discusses its relevance and importance within the forensic sciences.
Materials and methods
The collection of Diptera specimens was carried out in the following locations: Tres Cruces (22°55’06.01” S, 65°35’13.58” W), Humahuaca (23°12’14.27” S, 65°20’54.90” W), and Tumbaya (23°51’27.79” S, 68°28’03.31” W) (Fig. 1a-c). These locations are part of the Monte Province, Prepuna District in the province of Jujuy, Argentina. Two sampling campaigns were carried out, one during the dry season (June, July, and August) and the other during the wet season (December, January, and February) between 2016 and 2018. Each location was equipped with 18 traps, totaling 54 traps across all sampling locations, totaling 108 traps per year.
Figure 1. Location of the study area. The map depicts the region corresponding to the Monte Province, Prepuna District, in the province of Jujuy, Argentina. Study sites are located in the following localities: a) Tres Cruces, b) Humahuaca, and c) Tumbaya (Photos by Fernández Salinas, M. L.).
To obtain specimens in good condition for identification, traps were made following Hwang and Turner (2005) (Fig. 2). A modified cone trap, based on a soft drink bottle with a baited target, was constructed. The bottle traps were assembled using two 3 L clear plastic soft drink bottles with a diameter of 11.5 cm, along with a black acrylic container 11 cm measured in diameter and 13 cm in depth. Consisted of 2 parts, the upper collection chamber and the lower bait chamber. The collection chamber was formed from the bottles cut 24 cm and 12 cm from the top respectively, one pushed inside the other (so that the bottle’s spout acts as a funnel and prevents flies from escaping). The bait chamber was made with a black container so the flies were drawn upwards, into the transparent collection chamber. The 2 halves of the trap were push-fitted together and secured by strips of waterproof adhesive tape. To facilitate the entry of flies 4 holes of approximately 0.8 cm in diameter were made around the bait chamber. A 125 cm³ plastic container with the bait was placed at the base of this container. A feeding substrate made of 100 g of cow’s lung was used. A distance of approximately 100 m was maintained between the traps because of the competitive nature of the colonizing species and were separated from the floor as they were hung at a minimum height of 1.5 m to avoid the attack of scavenger mammals. It was placed in a closed recipient which was subjected to a warm temperature between 15 °C and 30 °C, during 60 hours, for sufficient time to decompose. The traps were left in place for 7 consecutive days.
The captured specimens were put in Kahn tubes with 70% alcohol and they were transported to the Institute of Altitude Biology (INBIAL), San Salvador de Jujuy, Jujuy, Argentina.
Figure 2. Design of the bottle necrotrap baited with cow lung.
Flies that presented mites were counted, separated by sex and identified to its most specific level using keys and revisions from Olea and Mariluis (2013), Whitworth (2014), and Mulieri et al. (2014). The flies were photographed “in situ” using a Canon 5D Mark IV camera, 3 extension tubes for macro photography and a Canon 85 mm 1.8 lens illuminated with a Godox AD200 flash and a Godox V860 flash. Afterward, the specimens were sent to the Parasitic Zoonoses Laboratory, National University of Mar del Plata, Mar del Plata, Buenos Aires, Argentina. Each fly was individually examined, and the number of mites per fly and their attachment sites on the host flies were determined. The mites were then removed with the assistance of fine-tipped needles. From selected mite specimens, permanent preparations in Hoyer’s medium were made. The remaining specimens were identified from temporary preparations after being cleared in lactic acid using an open slide technique in order to be observed under the optic microscope (Olympus CX31). Taxonomic identification was done at a genus level using diagnostic keys (Dindal, 1990). Mites were photographed with a Sony Powershot DSC-P200 camera. The photographs were edited with Adobe Photoshop CS.
Table 1
Percentage of prevalence and mean intensity of Myianoetus sp. associated with Compsomyiops fulvicrura, in 3 locations of Jujuy, Argentina.
Location | Season | Nº of flies with attached mites | Total Nº of flies | Nº of mites | Prevalence (%) | Mean Intensity |
Tumbaya | Dry | 26 | 661 | 243 | 3.93 | 9.34 |
Wet | 0 | 85 | 0 | 0 | 0 | |
Humahuaca | Dry | 40 | 1,517 | 478 | 2.63 | 11.95 |
Wet | 0 | 205 | 0 | 0 | 0 | |
Tres Cruces | Dry | 18 | 624 | 156 | 2.88 | 8.83 |
Wet | 10 | 582 | 276 | 1.71 | 27.3 |
The abundance of flies with attached mites was analyzed using Generalized Linear Models (GLM) through the software InfoStat (Di Rienzo et al., 2020). In the model, the 3 study locations were considered as fixed effects while the seasons were treated as random variables. Variance heterogeneity was adjusted using the VarExp variance function, and models were selected according to Akaike (AIC) and Bayesian (BIC) criteria. Subsequently, a Fisher’s LSD test (α = 0.05) of adjusted means and standard errors was conducted to evaluate differences between locations, following the methods described. Prevalence and mean intensity were calculated as indicated by Bush et al. (1997) and Margolis et al. (1982). Prevalence was calculated as the number of flies infected with phoretic mites, divided by the number of flies examined in a sample, and was expressed as a percentage. The mean intensity of phoresy was defined as the total number of phoretic mites of a particular species found in a sample, divided by the number of host flies.
For further taxonomic studies, voucher species were deposited as slide-mounted specimens in the Entomological Collection “Dr. Lilia Estela Neder”, Institute of Altitude Biology (INBIAL), National University of Jujuy, Jujuy, Argentina (INBIAL C 15000; INBIAL C 15001).
Results
A total of 9,454 Calliphoridae individuals were collected. They spanned 5 genera and 12 species: Calli-
phora vicina (Robineau-Desvoidy), Chlorobrachycoma versicolor (Bigot), Chrysomya albiceps (Wiedemann), Chrysomya megacephala (Fabricius), Cochliomyia mace-llaria (Fabricius), Cochliomya hominiborax (Coquerel), Compsomyiops fulvicrura (Robineau-Desvoidy), Com-psomyiops sp., Lucilia cuprina (Wiedemann), Lucilia sericata (Meigen), Sarconesia chlorogaster (Wiedemann), Sarconesiopsis magellanica (Le Guillou). The most abundant species were C. albiceps and C. fulvicrura with 4,651and 3,674 individuals respectively. C. fulvicrura was the only species that had mites attached to its body (Fig. 3). These mites primarily attached themselves to the thorax and head regions and were identified as deutonymphs of Myianoetus sp. (Figs. 4, 5, 6). The individuals found exhibit morphological similarities to the deutonymphs of Myianoetus muscarum (Linnaeus) (OConnor et al., 2015). However, they differ from this species by possessing dorsal hysterosomal setae of approximately equal length to the exobothridial setae, unlike M. muscarum, where the hysterosomal setae are less than half the length of the exobothridial setae. Given that this characteristic is diagnostic of M. muscarum, we hypothesize that the specimens uncovered in this study may represent a yet undescribed species. Out of the total number of C. fulvicrura individuals, 94 carried phoretic mites (83 females and 11 males), representing a prevalence of 2.56% (Table 1). A total of 1,153 mites were counted, which corresponds to a mean intensity of 12.26 mites/fly (1-89 rank) (Fig. 7). The majority of mites (76%) were found during the dry season in all 3 studied locations. However, in Tres Cruces, mites were also found during the wet season (Table 1).
Table 2
Summary of generalized linear model (GLM) analysis results and model fitting parameters. Significance levels (p values) and variance function parameters, model fitting parameters including number of observations (N), Akaike information criterion (AIC), Bayesian information criterion (BIC), the log probability, the standard deviation (Sigma) and the coefficient of determination (R2) are shown.
Effects | p-value | Variance function parameters |
Location | < 0.0001 ** | |
Season | 0.0105 * | -0.27 (dry) |
-0.09 (wet) |
Model tuning: N = 6, AIC = 31.83, BIC = 22.68, LogLik = -8.91, Sigma = 4.82, R2 = 0.74
Table 3
Results of the Fisher’s LSD test (α = 0.05): adjusted means and standard errors for the 3 locations under study. Common letters indicate that the means do not differ significantly (p > 0.05).
Location | Means | SE | |
Humahuaca | 20.84 | 1.98 | A |
Tumbaya | 6.84 | 1.98 | B |
Tres Cruces | 1.16 | 1.98 | C |
The GLM analysis revealed significant differences in the abundance of flies with attached mites among the study locations (p < 0.0001) and a significant effect of seasonality (p = 0.0105) (Table 2).
Figure 3. Deutonymphs of Myianoetus sp. (yellow arrow) between the thorax and abdomen of Compsomyiops fulvicrura.
Additionally, subsequent Fisher’s LSD analysis revealed statistically different groups among the study locations. A higher mean abundance of flies with attached mites was observed in Humahuaca, followed by Tumbaya and Tres Cruces (Table 3). It is noteworthy that the highest variance parameter for the dry season (-0.27) compared to the wet season (-0.09) suggests that these differences are primarily attributed to this time of the year, between June and August.
Discussion
The genus Myianoetus comprises more than 40 species widespread throughout the world (OConnor et al., 2015), most known only from deutonymphs phoretic on Diptera. In this work, the association between deutonymphs of Myianoetus sp. with C. fulvicrura is described for the first time. Up to present, there are reports of deutonymphs from the Myianoetus that have been found associated with various Diptera families: Sphaeroceridae (Fain et al., 1980), Muscidae (Centeno & Perotti, 1999; Greenberg & Carpenter 1960; Negm & Alatawi, 2011; Pimsler et al., 2016), Calliphoridae (Greenberg & Carpenter 1960; Miranda & Bermúdez, 2008) and Heleomyzidae (Zamec & Košel, 2014). Evidence obtained from lab experiments further described the phoretic interaction of the hypopi of Myianoetus muscarum with Muscina stabulans Fallen (Diptera: Muscidae), Stomoxys calcitrans Linnaeus (Diptera: Muscidae), Lucilia sericata (Diptera: Calliphoridae) and Musca domestica Linnaeus (Diptera: Muscidae) (Greenberg & Carpenter, 1960). Additionally, in a study carried out in Texas, USA, by Pimsler et al. (2016), a great number of M. muscarum individuals associated with Synthesiomyia nudiseta (Wulp) (Diptera: Muscidae) were collected in 3 human corpses.
Among the Calliphoridae species collected, C. albiceps stood out as the most abundant. However, deutonymphs of Myianoetus sp. were exclusively phoretically associated with C. fulvicrura. The statistical differences observed in the abundances of flies with attached mites among the different studied locations suggest that these were influenced by the dry season. Therefore, the preference for C. fulvicrura could be associated with seasonal variation, as it was more abundant during the dry season, contrasting with C. albiceps, which showed a preference for the wet season. These trends were notable in Tumbaya and Humahuaca, where C. albiceps was the dominant species, while in Tres Cruces, the abundance of this species was very low, with C. fulvicrura being the dominant species in both seasons in that area. Additionally, it is plausible that this choice is related to the chemical attraction of mites to volatile substances released by the puparia of C. fulvicrura, as demonstrated in the studies by Greenberg and Carpenter (1960). These observations were reflected in the prevalence values, which indicated higher values during the dry season in all 3 locations, compared to the wet season.
Figure 5. Dorsal view of Myianoetus sp. (scale = 0.05 mm). The yellow arrow indicates the hysterosomal setae (ex) and exobothridial setae (in).
Figure 6. Dorsal view of legs I and II of Myianoetus sp. (scale = 0.1 mm). The yellow arrow shows the bifurcate empodial claw, characteristic of the genus, present on legs I-III.
Figure 7. Abundance frequency (AF) histogram of Myianoetus sp. deutonymphs associated with Compsomyiops fulvicrura individuals.
The lack of interaction between Myianoetus sp. with other species and its demonstrated affinity with C. fulvicrura suggest that these mites can be phoretically selective in the Prepuna environment. At genus or species level, mites have micro-habitat specific requirements, being excellent specific environmental indicators, offering themselves as potentially one of the most informative pieces of biological trace evidence collected from a crime scene (Perotti & Braig, 2019). This may explain events of corpse location, of relocation, a link to a suspect and a possible connection between a suspect and a victim or a crime scene (Hani et al., 2018; Kamaruzaman et al., 2018; Szeleczl et al., 2018). The specificity and abundance of mites, coupled with the intensity of phoresy, could contribute to estimating more precise post-mortem intervals (PMI) (Miranda & Bermúdez, 2008; Rodrigueiro & Prado, 2004; Russell et al., 2004). In addition, Perotti and Braig (2009) suggested that the presence of a specific phoretic mite (for example Myianoetus sp.) may confirm the presence of its host (for example C. fulvicrura), even when the host is already gone.
Given that mites are a valuable forensic tool, it is crucial to deepen the understanding of the biology and ecology of the species involved. To expand this knowledge, it is necessary to continue registering and investigating new species and their phoretic associations under various climatic and biogeographical conditions.
Acknowledgements
We would like to thank Pablo A. Martínez for his critical analysis and recommendations for our manuscript, Mario A. Linares for the Myianoetus sp.photograph and Ismael Acosta for the C. fulvicrura photograph.
References
Arnaldos, M. I., García, M. D., Romera, E., Presa, J. J., & Luna, A. (2005). Estimation of post-mortem interval in real cases based on experimentally obtained entomological evidence. Forensic Science International, 149, 57–65. https://doi.org/10.1016/j.forsciint.2004.04.087
Barton, P. S., Weaver, H. J., & Manning, A. D. (2014). Contrasting diversity dynamics of phoretic mites and beetles associated with vertebrate carrion. Experimental and Applied Acarology, 63, 1–13. https://doi.org/10.1007/s10493-013-9758-7
Braig, H. R., & Perotti, M. A. (2009). Carcasses and mites. Experimental and Applied Acarology, 49, 45–84. https://doi.org/10.1007/s10493-009-9287-6
Bush, A. O., Lafferty, K. D., Lotz, J. M., & Shostak, A. W. (1997). Parasitology Meets Ecology on Its Own Terms: Margolis et al. Revisited. The Journal of Parasitology, 83, 575–583. https://doi.org/10.2307/3284227
Camerik, A. M. (2010). Phoresy revisited. In M. Sabelis, & J. Bruin (Eds.), Trends in Acarology (pp. 333–336). Dordrecht: Springer. https://doi.org/10.1007/978-90-481-9837-5_53
Catts, E. P., & Goff, M. L. (1992). Forensic entomology in criminal investigations. Annual Review of Entomology, 37, 253–272. https://doi.org/10.1146/annurev.en.37.010192.001345
Centeno, N. D., & Perotti, M. A. (1999). Ácaros vinculados a procesos de descomposición de cadáveres y sus posibles asociaciones foréticas. In Actas y Trabajos de la XIX Reunión Argentina de Ecología, 1999. Tucumán, Argentina.
Di Rienzo J. A., Casanoves, F., Balzarini, M. G., González, L., Tablada, M., & Robledo, C. W. (2020). InfoStat versión 2020. Centro de Transferencia InfoStat, FCA, Universidad Nacional de Córdoba, Argentina.
Dindal, D. L. (1990). Soil biology guide. New York: John Wiley and Sons.
Fain, A., Britt, D. P., & Molyneux, D. H. (1980). Myianoetus copromyzae sp. nov. (Acari, Astigmata, Anoetidae) phoretic on Copromyza atra (Meigen 1830) in Scotland. Journal of Natural History, 14, 401–403. https://doi.org/
10.1080/00222938000770341
Farish, D. J., & Axtell, R. C. (1971). Phoresy redefined and examined in Macrocheles muscaedomesticae (Acarina: Macrochelidae). Acarologia, 13, 16–29.
Greenberg, B., & Carpenter, P. D. (1960). Factors in phoretic association of a mite and fly. Science, 132, 738–739. https://doi.org/10.1126/science.132.3429.738
Hani, M., Thieven, U., & Perotti, M. A. (2018). Soil bulb mites as trace evidence for the location of buried money. Forensic Science International, 292, e25–e30. https://doi.org/10.1016/j.forsciint.2018.09.016
Heo, C. C., Teel, P. D., & OConnor, B. M. (2021). Acari community in association with delayed pig carrion decomposition. Experimental and Applied Acarology, 85, 223–246. https://doi.org/10.1007/s10493-021-00676-6
Hwang, C., & Turner, B. D. (2005). Spatial and temporal variability of necrophagous Diptera from urban to rural areas. Medical and Veterinary Entomology, 19, 379–391. https://doi.org/10.1111/j.1365-2915.2005.00583.x
Kamaruzaman, N. A. C., Mašán, P., Velásquez, Y., González-Medina, A., Lindström, A., Braig, H. R. et al. (2018). Macrocheles species (Acari: Macrochelidae) associated with human corpses in Europe. Experimental and Applied Acarology, 76, 453–471. https://doi.org/10.1007/s10493-018-0321-4
Margolis, L., Esch, G. W., Holmes, J. C., Kuris, A. M., & Schad, G. A. (1982). The use of ecological terms in parasitology (report of an ad hoc committee of the American Society of Parasitologists). Journal of Parasitology, 68, 131–133. https://doi.org/10.2307/3281335
Miranda, R., & Bermúdez, S. (2008). Ácaros (Arachnida: Acari) asociados con moscas Calliphoridae (Diptera: Oestroidea) en tres localidades de Panamá. Revista Colombiana de Entomología, 34, 192–196. https://doi.org/10.25100/socolen.v34i2.9287
Mulieri, P. R., Mariluis, J. C., & Patitucci, L. D. (2014). Calliphoridae. In S. Roig-Juñent, L. E. Claps, & J. J. Morrone (Eds.), Biodiversidad de artrópodos argentinos, Vol. 4 (pp. 463–474). INSUE, Universidad Nacional de Tucumán, San Miguel de Tucumán, Argentina.
Negm, M. W., & Alatawi, F. J. (2011). Four new records of mites (Acari: Astigmata) phoretic on insects in Riyadh, Saudi Arabia. Journal of the Saudi Society of Agricultural Sciences, 10, 95–99. https://doi.org/10.1016/j.jssas.2011.04.001
OConnor, B. M. (1982). Evolutionary ecology of astigmatid mites. Annual Review of Entomology, 27, 385–409. https://doi.org/10.1146/annurev.en.27.010182.002125
OConnor, B. M. (2009). Astigmatid mites (Acari: Sarcoptiformes) of forensic interest. Experimental and Applied Acarology, 49, 125–133. https://doi.org/10.1007/s10493-009-9270-2
OConnor, B. M., Pimsler, M. L., Owings, C. G., & Tomberlin, J. K. (2015). Redescription of Myianoetus muscarum (Acari: Histiostomatidae) associated with human remains in Texas, USA, with designation of a neotype from Western Europe. Journal of Medical. Entomology, 52, 539–550. https://doi.org/10.1093/jme/tjv045
Olea, S. M., & Mariluis, J. C. (2013). The genus Calliphora (Diptera: Calliphoridae) in Argentina, with the first records of C. lopesi Mello 1962. Revista de la Sociedad Entomológica Argentina, 72, 99–104.
Perotti, M. A., & Braig, H. R. (2009). Phoretic mites associated with animal and human decomposition. Experimental and Applied Acarology, 49, 85–124. https://doi.org/10.1007/s10493-009-9280-0
Perotti, M. A., Braig, H. R., & Goff, M. L. (2010). Phoretic mites and carcasses: Acari transported by organisms associated with animal and human decomposition. In J. Amendt, M. Goff, C. Campobasso, & M. Grassberger (Eds.), Current concepts in forensic Entomology (pp. 69–91). Dordrecht: Springer. https://doi.org/10.1007/978-1-4020-9684-6_5
Perotti, M. A., & Braig, H. R. (2019). Acarology in Crimino-Legal Investigations. In J. Byrd, & J. Tomberlin (Eds.), Forensic Entomology, 3rd. Edition (pp. 461–473). Boca Raton: CRC Press. https://doi.org/10.4324/9781351163767-22
Pimsler, M. L., Owings, C. G., Sanford, M. R., OConnor, B. M., Teel, P. D., Mohr, R. M., & Tomberlin, J. K. (2016). Associa-
tion of Myianoetus muscarum (Acari: Histiostomatidae) with Synthesiomyia nudiseta (Wulp) (Diptera: Muscidae) on human remains. Journal of Medical Entomology, 53, 290–295. https://doi.org/10.1093/jme/tjv203
Rai, J., Amendt, J., Bernhardt, V., Pasquerault, T., Lindström, A., & Perotti, M. A. (2020). Mites (Acari) as a relevant tool in trace evidence and postmortem analyses of buried corpses. Journal of Forensic Sciences, 65, 2174–2183. https://doi.org/10.1111/1556-4029.14506
Rodrigueiro, T. S. C., & Prado, A. P. (2004). Macrocheles muscaedomesticae (Acari, Macrochelidae) and a species of Uroseius (Acari, Polyaspididae) phoretic on Musca domestica (Diptera, Muscidae): effects on dispersal and colonization of poultry manure. Iheringia. Série Zoologia, 94, 181–185. https://doi.org/10.1590/S0073-47212004000200011
Russell, D. J., Schulz, M. M., & OConnor, B. M. (2004). Mass occurrence of astigmatid mites on human remains. Abhandlungen und Berichte des Naturkundmuseums Görlitz, 76, 51–56.
Saloña-Bordas, M. I., & Perotti, M. A. (2015). Acarología forense. Ciencia Forense, 12, 91–112.
Szelecz, I., Lösch, S., Seppey, C. V. W., Lara, E., Singer, D., Sorge, F. et al. (2018). Comparative analysis of bones, mites, soil chemistry, nematodes and soil micro-eukaryotes from a suspected homicide to estimate the post-mortem interval. Scientific Reports, 8, 25. https://doi.org/10.1038/s41598-017-18179-z
Whitworth, T. (2014). A revision of the Neotropical species of Lucilia Robineau- Desvoidy (Diptera: Calliphoridae). Zootaxa, 3810, 1–76. https://doi.org/10.11646/zootaxa.3810.
1.1
Zamec, R., & Košel, V. (2014). A new species of mite (Acari: Histiostomatidae) phoretic on Gymnomus caesius (Diptera: Heleomyzidae) from Vlčie Diery cave. Biologia, 69, 916–919. https://doi.org/10.2478/s11756-014-0387-3
Ringtails (Bassariscus astutus) as seed dispersers in an urban gradient under conditions of low human activity due to COVID-19
Joselin Judith Peña-Herrera a, Yury Glebskiy a, b, *, Teresa de Jesús Hernández-Trejo a, Zenón Cano-Santana a
a Universidad Nacional Autónoma de México, Facultad de Ciencias, Departamento de Ecología y Recursos Naturales, Laboratorio de Interacciones y Procesos Ecológicos, Circuito Exterior s/n, Ciudad Universitaria, Coyoacán, 04510 Ciudad de México, Mexico
b Universidad Nacional Autónoma de México, Facultad de Ciencias, Posgrado en Ciencias Biológicas, Circuito Exterior s/n, Ciudad Universitaria, Coyoacán, 04510 Ciudad de México, Mexico
*Corresponding author: agloti@ciencias.unam.mx (Y. Glebskiy)
Received: 14 September 2023; accepted: 6 March 2024
Abstract
Seed dispersal by animals is a key ecosystemic process in many environments; however, it could be compromised or increased in urban environments due to changes in the landscape, the introduction of exotic species, and human activities. This article aims to evaluate the role of ringtails (Bassariscus astutus) as seed dispersers in an urban-natural gradient during low human activity due to the COVID-19 pandemic. Ringtail feces were collected in 3 sampling sites with different levels of urbanization (ranging from 100 to 5% of natural vegetation), and the seeds germinated in germination chambers. Twenty species of plants were dispersed by ringtails, more than reported in previous studies. More seeds were dispersed in natural (7.1 seeds per g) than urbanized (3.2 seeds per g) areas, but diversity and richness were higher in urbanized areas. This suggests that urban environments have a greater diversity, and it could be attributed to the microenvironments created by urban infrastructure and the exotic plants that are established in the area.
Keywords: Endozoochory; Mexico City; Opuntia; REPSA; Zoochory
© 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/).
Cacomixtles (Bassariscus astutus) como dispersores de semillas en un gradiente urbano bajo condiciones de baja actividad humana por COVID-19
Resumen
La dispersión de semillas por animales es un proceso clave en muchos ecosistemas, pero este se puede ver comprometido o incrementado en ambientes urbanos debido a cambios en el paisaje, introducción de especies exóticas y actividades humanas. El objetivo de este artículo es evaluar el papel del cacomixtle (Bassariscus astutus) como dispersor de semillas en un gradiente urbano-natural durante un periodo de baja actividad humana debido a la pandemia de COVID-19. Se colectaron excretas de cacomixtles en 3 localidades con diferente grado de urbanización (entre 100 y 5% de vegetación natural) y las semillas fueron germinadas en cámaras de germinación. Se registraron 20 especies de plantas dispersadas, más que lo reportado en estudios previos. Más semillas fueron dispersadas en áreas naturales (7.1 semillas por g) que urbanizadas (3.2 semillas por g), pero la riqueza y diversidad fueron mayores en áreas urbanizadas. Ésto sugiere que la diversidad en ambientes urbanos es mayor, lo cual se puede atribuir a los microambientes formados por la infraestructura urbana y las plantas exóticas establecidas en el área.
Palabras clave: Endozoocoria; Ciudad de México; Opuntia; REPSA; Zoocoria
Introduction
Some animals may provide a key ecosystemic service, acting as seed dispersers, which allow many plant species to effectively move their offspring and maintain plant communities across different environments. An example of those animals is the ringtail (Procyonidae: Bassariscus astutus) which is an omnivorous and opportunistic animal that has shown potential to disperse the seeds of a great number of plant species across many ecosystems (Alexander et al., 1994; Rodríguez-Estrella et al., 2000; Rubalcava-Castillo et al., 2020). However, this important service could be diminished in urban ecosystems, where ringtails are very common (Barja & List, 2006; Swanson et al., 2022). Because urban areas offer new sources of food like anthropogenic waste and a variety of exotic plants, ringtails could disperse fewer seeds or seeds of exotic plants. This is an important concern since urban ecosystems are growing fast and thus becoming important areas for the conservation of species and areas on which we rely to obtain ecosystemic services.
Previous studies suggest that seed dispersal is diminished in cities due to a great number of unsuitable habitats (Cheptou et al., 2008) and barriers that obstruct animal movement and thus seed dispersal (Niu et al., 2018). However, animal-plant networks are persistent in cities (Cruz et al., 2013), and plants that rely on animal dispersal tend to have a more successful regeneration of populations than plants that rely on other strategies for dispersal (Niu et al., 2023). At the same time, most studies show that urban ecosystems tend to be more diverse than natural areas due to the great number of exotic species, microenvironments that can host greater plant diversity, and the fact that cities tend to be built in highly diverse locations (Kühn et al., 2004; Wania et al., 2006). Yet all these studies are performed in urban areas with both human activity and urban infrastructure; therefore, the hypothesis that urban areas are diverse due to the exotic plants and microhabitats and not due to “direct-human” seed dispersal (for example, seeds we throw away as garbage) is yet to be proven.
Particularly for the ringtails, Cisneros-Moreno and Martínez-Coronel (2019) found differences in the urban and rural ringtail diets. They report that in urban environments, ringtails consume 11 plant species, and 9 in rural environments. Other studies also show that ringtails commonly consume human-generated waste from trash cans when it is available (Castellanos et al., 2009; Picazo & García-Collazo, 2019). However, all those studies were made under normal human activity, but if it is reduced, the generation of waste could be diminished, affecting the ringtail diet, which in turn could lead to a change in their role as seed dispersers.
Therefore, this article aims to compare the diversity, abundance, and species of seeds dispersed by ringtails in 3 areas with different levels of urbanization during a time of reduced human activity due to the COVID-19 pandemic.
Materials and methods
This study was performed inside the main campus of the Universidad Nacional Autónoma de Mexico, located in Mexico City, Mexico. The campus contains a well-preserved ecological reserve (Reserva Ecológica del Pedregal de San Ángel; henceforth REPSA) and urban areas such as buildings and roads, all of which are surrounded by the Mexico City (Fig. 1). Therefore, a gradient between urban and natural areas can be found in a relatively small area (730 ha; Zambrano et al., 2016) that otherwise shares all environmental characteristics like mean temperature (18.2°C), precipitation (752 mm), original substrate, and vegetation: xerophitic shrubs (Rzedowski, 1954; SMN, 2023).
An important characteristic of this study is that it was performed under lockdown conditions due to the COVID-19 pandemic. Because of the rapid spreading of the virus, most activities on campus were switched to virtual mode; students took lessons from home; maintenance such as cleaning, gardening, and security were kept to a bare minimum; and all research activities had to be done from home, except for some specific cases, such as this study, that required special permission from the university. Those measures were implemented in March 2020 and began to be gradually lifted in the spring of 2022. Under normal conditions, the campus is visited by 166,474 people, and 70,000 vehicles, and 15 tons of waste (excluding gardening products) are generated (Zambrano et al., 2016). However, as a result of the pandemic, human activity on campus such as driving, waste generation and gardening, among others was minimal for 2 years. This gave us the opportunity to study the ringtail seed dispersal in an urban environment without human presence.
The correct identification of ringtail feces is an essential part of this project, since, in our study site, they could be confused with excretes of opossums (Didelphis virginiana). Previous studies suggest that opossums do not use latrines (Aranda, 2000), however, to ensure that latrines are used exclusively by ringtails, we placed camera traps in front of 8 latrines, and animal interactions with those latrines were recorded.
Figure 1. Map of the study location. Blue dots, Faculty (most urbanized area) latrines; red dots, the Institute latrines (semi-urban area); and green dots, the West core (natural area). The lines represent the 178 m around the latrines in which the percentages of the different types of terrain were calculated.
Three areas separated by at least 800 m were chosen to represent the urban-natural gradient: natural, semi-urban, and urban (Fig. 1). The home ranges of this species in the location are small: between 3 and 9.9 ha (Castellanos & List, 2005); therefore, this separation should ensure independence between the treatments. The level of urbanization was based on the percentage of area covered by natural vegetation, altered vegetation, impermeable areas without buildings (mainly roads and parking lots), and buildings found in a 178 m area (Fig. 1, Table 1) around the sampling points (178 m is the radius of the maximum activity area reported for ringtails in this particular location; Castellanos & List, 2005). The natural area (henceforth the West core) is located inside the west core of the REPSA. Vegetation consists mainly of shrubs, and Opuntia cactus is quite common (Cano-Santana, 1994). The semi-urban area (henceforth Institutes) is located around the humanitarian institutes area and consists of a mosaic of spatially located buildings, parking lots, altered vegetation (grass and some cultivated trees, mostly without fruit) and remnants of natural vegetation that surround this area on all sides (Fig. 1, Flores-Morales, 2023). The urban area (henceforth Faculty) is located inside the faculty of sciences and is dominated by tightly packed buildings divided by impermeable areas and gardens. The vegetation is diverse and includes a small amount of native plants such as Opuntia, but mostly consists of grass and introduced trees some of which have fruits that could be consumed by ringtails (Mendoza-Hernández & Cano-Santana, 2009). During the lockdown, gardens were left mostly unattended, and some alimentary plants began to grow (for example, we encountered several tomato plants with ripe fruits). The natural vegetation is located mostly on the edges of this area (Fig. 1).
Table 1
Amount of terrain coverage in the sampling areas. Impermeable areas were considered all areas covered with concrete but without buildings, mainly roads and parking lots.
Natural vegetation (%) | Altered vegetation (%) | Buildings (%) | Impermeable areas (%) | Total area (ha) | |
West core | 100 | 0 | 0 | 0 | 34.1 ha |
Institutes | 49.2 | 24.2 | 7 | 19.6 | 30.7 ha |
Faculty | 4.9 | 35 | 30.5 | 29.6 | 15.4 ha |
To estimate the number of seeds dispersed by ringtails, we collected excretes from 13 latrines in each location, on January 18th 2022. Since previous studies report that seeds in our study area commonly have dormancy (Glebskiy, 2019), half of each latrine excretes was placed in plastic bags and half in mesh bags (with 1 × 1 mm openings). Both treatments were left in the field (on the soil and without cover) until April 29th 2022 (when the rains and thus the germination period started). The advantage of the plastic bag treatment is that it allows the seeds to experience the temperature changes that are responsible for ending dormancy in most plants and protects the seeds, but limits some other factors like gas exchange and humidity that could potentially contribute to the end of dormancy (Baskin & Baskin, 1998). On the other hand, the mesh bag allows for a better representation of the natural conditions to which seeds are exposed but is susceptible to losing seeds through the mesh and the addition of new seeds from the environment. Both methods were tested.
At the beginning of the rainy season (April 29th 2022) bags were collected and put to germinate in commercial soil (the soil was sterilized by microwaving, and a control with no excretes was used to test the efficiency of sterilization and possible future seed additions) in a germination chamber (25°C, 16 hours of light, and 8 hours of darkness) for 4 months. All plant germinations were recorded and identified to morphospecies; when the plants grew, they were identified to the finest level possible. Species that germinated in the control pot were considered later additions and excluded from the analysis.
Data were analyzed with the R statistical packages: stats (R core team, 2022), dunn.test (Dinno, 2017), fossil (Vavrec, 2011), and vegan (Oksanen et al., 2022). Number of germinated seeds per gram of excrete was calculated. To compare between treatments (3 locations and 2 types of protection bags) Kruskal-Wallis and Dunn tests were performed for the total number of germinated seeds and the number of Opuntia seeds (this was the only species with enough data to analyze independently). A GLIM test (with Poisson distribution) was used to determine if the level of protection (plastic bags = 1, mesh bags = 0) and amount of vegetation (both natural and altered) influenced the amount of total and Opuntia germinated seeds. To compare the richness and diversity of dispersed seeds, we calculated the Shannon diversity and Chao 1 (± 95% confidence interval) richness. The similarity between treatments was measured using the Jaccard and Bray-Curtis indexes.
Results
A total effort of 329 trap nights was performed with 199 independent records (at least 1 hour between sightings) of ringtails, of which 44 times ringtails defecated in the latrine (Fig. 2). Opossums were observed 66 times, and no defecation was observed. At the same time, there were 9 observations of rodents and 7 of birds feeding from the latrines.
A total of 52 bags were recovered from the field (34 plastic, and 18 mesh bags), and were put to germinate. In total, 782 plants of 20 species (Table 2) germinated in this experiment. More seeds per gram of excrete germinated in the mesh bags treatment (U tests; V = 1,225, p < 0.001); however, more Opuntia seeds germinated in the plastic bag treatment (U test; V = 300, p < 0.001; Table 3). The Kruskal-Wallis test showed no differences in the number of germinated seeds between locations in the mesh bags, but there were significant differences for total seed and Opuntia seeds in the plastic bag treatment (p = 0.042 and p = 0.002). According to the Dunn test, there were fewer total seeds in the Faculty area than the West core (p = 0.009) and fewer Opuntia seeds in the Faculty area than the Institutes (p = 0.015) and the West core (p = 0.009).
GLIM analysis showed that both level of protection (-0.818, p < 0.001) and vegetation area (1.279, p < 0.001) are significant predictors for the number of total seeds germinated. For Opuntia seeds, protection (1.674, p < 0.001) and vegetation (3.559, p < 0.001).
Figure 2. a) Ringtail defecating in a latrine; b) a rodent consuming seeds from a latrine; c) a latrine; d) plants germinated from excretes.
We found significant differences in richness (according to Chao1) between the Faculty area and the West core (in plastic bags) and between the Faculty area and Institutes and the West core (in mesh bags; Table 4). The Bray-Curtis dendrogram shows 2 important groups: the mesh bag treatment and the plastic bag (Fig. 3). Jaccard index showed the following results within the same location,Faculty plastic-Faculty mesh: 0.714, Institutes plastic-Institutes mesh: 0.546, West core plastic-West core mesh: 0.571, total plastic-total mesh: 0.65; between locations,plastic bags, Faculty plastic-Institutes plastic: 0.769, Faculty plastic-West core plastic: 0.5, Institutes plastic-West core plastic: 0.643;between locations, mesh bags, Faculty mesh-Institutes mesh: 0.5, Faculty mesh-West core mesh: 0.294, Institutes mesh-West core mesh: 0.333.
Figure 3. Bray-Curtis dendrogram of germination treatments. First letter represents the location, W: west core, I: Institutes, F: Faculty; second letter, the type of bag: M: mesh, P: plastic.
Discussion
The camera trap experiment was designed to prove that the latrines from which we collected excretes belong to ringtails and not opossums since those animals produce very similar feces (Aranda, 2000). Given that we observed 44 events of defecation by ringtails and zero by opossums, it we can conclude that ringtails are the only latrine users. However, at the same time, rodents and birds were seen feeding in those latrines (Fig. 2). The ratio for those visits is 1 visit per 2.8 defecations, and this is important for seed dispersal, since most likely those visitors feed on seeds that are dropped by ringtails and can selectively remove some species. At the same time, it is interesting to consider the trade-off for the rodents that consume seeds from ringtail latrines since they are exposed to predation by the latrine owners. Although it is outside the scope of this research, we consider that the interaction around the latrines is a topic that should be further investigated to better understand the role of these animals as seed dispersers and the interactions that the latrines produce.
The comparison between plastic and mesh bags shows differences between the treatments, especially in the number of seeds that germinated per gram of excrete (Tables 2, 3). This could be attributed to the fact that mesh bags allow the smallest particles to exit the bag, and the total weight of feces diminishes while the number of seeds remains constant. Therefore, the mesh bag treatment is not adequate for estimating the number of seeds per gram of excrete, although data concerning the diversity and richness of seeds is still valid. Given the above and the fact that there were more plastic bag replicas, the data interpretation in this research is based on the plastic bag treatment.
Table 2
Number of seeds per gram of excrete in all treatments for each plant species.
Species | Common name | Faculty | Institutes | West core | |||
Plastic | Mesh | Plastic | Mesh | Plastic | Mesh | ||
Ageratina sp. 1 | Snakeroot | 0.11 | 3.19 | 0.41 | 2.95 | 0.12 | 8.63 |
Ageratina sp. 2 | Snakeroot | 0.02 | 0.09 | 0.02 | 0 | 0 | 0 |
Asclepias linaria | Pineneedle milkweed | 0 | 0 | 0 | 0 | 0.13 | 0 |
Amaranthaceae | 0.33 | 0.19 | 0 | 0 | 0 | 0 | |
Bidens sp. | Beggarticks | 0 | 0 | 0.02 | 0 | 0.05 | 0.61 |
Cissus sicyoides | Princess vine | 0 | 0 | 0 | 0 | 0.03 | 0.33 |
Conyza sp. | Horseweed | 0.07 | 0.22 | 0.26 | 0 | 0 | 0 |
Drymaria laxiflora | Chickweed | 0.04 | 0.27 | 0.12 | 0.35 | 0.03 | 2.68 |
Eragrostis sp. | Lovegrass | 0 | 0.09 | 0 | 0 | 0 | 0 |
Iresine sp. | Bloodleaf | 0 | 0 | 0 | 0 | 0 | 2.04 |
Opuntia sp. | Prickly pear | 0.23 | 0.98 | 2.23 | 0.59 | 6.36 | 0.24 |
Solanaceae | 0.03 | 0.27 | 0.08 | 0 | 0.1 | 0.71 | |
Solanum nigrescens | Slender nightshade | 0.1 | 2.79 | 0.04 | 2.14 | 0.08 | 0 |
Stevia sp. | Stevia | 0 | 0.09 | 0 | 0 | 0 | 0 |
Phytolacca icosandra | Tropical pokeweed | 0.46 | 0.15 | 0.25 | 0.22 | 0.08 | 0 |
Poaceae 1 | 0.02 | 0.39 | 0.12 | 0.07 | 0.13 | 0.17 | |
Poaceae 2 | 0 | 0 | 0 | 0 | 0 | 0.25 | |
Unknown sp. 1 | 0.02 | 0 | 0.1 | 0 | 0.05 | 0.33 | |
Unknown sp. 2 | 0 | 0 | 0 | 0 | 0.05 | 0 | |
Unknown sp. 3 | 1.72 | 0 | 0 | 0 | 0 | 0 |
Table 3.
Number of Opuntia and total seeds dispersed by ringtails in different environments. Values given in seeds per g of excrete ± S.D.
Total seeds | Opuntia seeds | |||
Plastic | Mesh | Plastic | Mesh | |
West core | 7.13 ± 8.06 | 15.99 ± 13.85 | 6.36 ± 7.58 | 0.24 ± 0.3 |
Institutes | 3.64 ± 2.87 | 6.33 ± 6.61 | 2.23 ± 2.72 | 0.59 ± 1.33 |
Faculty | 3.16 ± 7.09 | 8.72 ± 5.51 | 0.23 ± 0.75 | 0.98 ± 2.39 |
Total | 4.7 ± 6.64 | 10.88 ± 10.26 | 2.96 ± 5.46 | 0.58 ± 1.5 |
GLIM analysis shows a positive relation between the number of seeds dispersed by ringtails and vegetation cover. Therefore, ringtails disperse fewer seeds in urban areas (Table 3); however, diversity and richness tend to be higher in the Faculty, the most urbanized location (Table 4).
Table 4
Richness (Chao1 ± 95% confidence intervals) and diversity (Shannon) of seeds dispersed by ringtails.
Faculty | Institutes | West core | Total | |||||||||
Plastic | Mesh | Total | Plastic | Mesh | Total | Plastic | Mesh | Total | Plastic | Mesh | Total | |
Observed richness | 12 | 12 | 14 | 11 | 6 | 11 | 12 | 10 | 14 | 16 | 17 | 20 |
Chao1 | 16 ± 2.65 | 14.67 ± 1.85 | 18.5 ± 3.9 | 15 ± 2.65 | 8 ± 1.87 | 17 ± 0 | 12.17 ± 0.34 | 10.25 ± 0.44 | 14.5 ± 0.73 | 16.25 ± 0.44 | 19.67 ± 1.85 | 24.5 ± 3.4 |
Shannon | 1.45 | 1.839 | 1.889 | 1.273 | 1.434 | 1.466 | 0.564 | 1.758 | 1.122 | 1.403 | 2.041 | 1.779 |
This is consistent with most previous studies that show a greater diversity of seeds in urban settings (Kühn et al., 2004; Wania et al., 2005), and that ringtails consume more plant species in the cities (Cisneros-Moreno & Martínez-Coronel, 2019). At our location, ringtails disperse fewer seeds in urban areas, but their richness and diversity are higher. The novel contribution of this research is that it was performed in an urban area with very little human activity; therefore, the direct effect of human activities such as waste generation and gardening could not explain the differences between treatments. At the same time, all 3 locations share very similar climatic conditions and originally were the same ecosystem. This suggests that differences between treatments are not due to original conditions (although it has been proven in other locations; Kühn et al., 2004). The high diversity of seeds in urbanized areas can be attributed to the great variety of habitats found in an urban area and the exotic species that have already been established there.
Ringtails are common in urban areas, and Castellanos et al. (2009) suggest they prefer disturbed areas over natural ones. This could be due to food availability in urbanized areas, and previous studies show that ringtails have a greater variety of food in urban areas. For example, Cisneros-Moreno and Martínez-Coronel (2019) show that an urban population consumes 36 food items and a rural population registered 28 items. Another advantage that cities offer is infrastructure. Poglayen-Neuwall and Toweill (1988) mention that rocky habitats favor ringtails, and human-made buildings provide a similar habitat. Interestingly, some data suggests that in the absence of humans, sightings of those animals in urban areas increased (Tzintzun-Sánchez, 2022; pers. obs.). This suggests that the benefits ringtails obtain from cities don’t come directly from human activities since, even without our presence, those animals have the benefits of infrastructure and a more diverse diet. However, further studies are needed to fully prove this hypothesis.
Previous studies have reported that fruits are important in the diet of B. astutus; however, the number of different plant species is generally low (13, ~ 9, 10, Alexander et al. [1994], Harrison [2012], and Rodríguez-Estrella et al. [2000], accordingly). In this study, 20 species of seeds were found and it is noteworthy that the cited articles included all vegetal matter in their analysis while this study only considers seeds that germinated. This suggests that the diversity of plants that are consumed in our site is much higher than usual. This could be the result of higher plant diversity in the location, or that ringtails tend to have a more frugivore diet here, this last hypothesis is supported by personal observations of the authors.
Based on the above, it could be concluded that there are several animals that consume the latrine contents, and this interaction could be important for seed dispersal. Ringtails disperse more seeds in natural areas, but their richness is higher in urban areas even in the absence of human activity.
Acknowledgements
We are thankful to M. E. Muñiz-Díaz for facilitating the germination chambers and her advice during this research, to Y. Martínez-Orea for helping with plant identification, to I. Castellanos-Vargas for technical support, and to the SEREPSA working team for facilitating the permission to perform this research. This project was financially supported by PAPIIT grant IN212121 (“El efecto de la urbanización sobre el tlacuache Didelphis virginiana en un matorral xerófilo de la Ciudad de México”), awarded to ZCS.
References
Alexander, L. F., Verts, B. J., & Farrell, T. P. (1994). Diet of ringtails (Bassariscus astutus) in Oregon. Northwestern Naturalist, 75, 97–101. https://doi.org/10.2307/353683131
Aranda, M. (2000). Huellas y otros rastros de los mamíferos grandes y medianos de México. Cuernavaca, Morelos: Instituto de Ecología, A.C./ la Comisión Nacional para el conocimiento y Uso de la Biodiversidad.
Barja, I., & List, R. (2006). Faecal marking behaviour in ringtails (Bassariscus astutus) during the non-breeding period: spatial characteristics of latrines and single faeces. Chemoecology, 16, 219–222. https://doi.org/10.1007/s00049-
006-0352-x
Baskin, C. C., & Baskin, J. M. (1998). Seeds: ecology, biogeography, and, evolution of dormancy and germination. San Diego: Academic Press.
Cano-Santana, Z. (1994). Flujo de energía a través de Sphenarium purpuracens (Orthoptera: Acrididae) y productividad primaria neta aérea en una comunidad xerófila (Ph.D. Thesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. México D.F.
Castellanos, G., García, N., & List, R. (2009). Ecología del cacomixtle (Bassariscus astutus) y la zorra gris (Urocyon cinereoargenteus). In A. Lot, & Z. Cano (Eds.), Biodiversidad del ecosistema del Pedregal de San Ángel (pp. 371–381). Ciudad de Mexico: Secretaria Ejecutiva de la REPSA-UNAM.
Castellanos, G., & List, R. (2005). Área de actividad y uso de hábitat del cacomixtle (Bassariscus astutus) en “El Pedregal de San Ángel”. Revista Mexicana de Mastozoología, 9, 113–122.
Cheptou, P. O., Carrue, O., Rouifed, S., & Cantarel, A. (2008). Rapid evolution of seed dispersal in an urban environment in the weed Crepis sancta. Proceedings of the National Academy of Sciences, 105, 3796–3799. https://doi.org/10.1073/pnas.0708446105
Cisneros-Moreno, C., & Martínez-Coronel, M. (2019). Alimentación del cacomixtle (Bassariscus astutus) en un ambiente urbano y uno agrícola en los valles centrales de Oaxaca. Revista Mexicana de Mastozoología (Nueva Época), 9, 31–43. https://doi.org/10.22201/ie.20074484e.2019.1.1.274
Cruz, J. C., Ramos, J. A., Da Silva, L. P., Tenreiro, P. Q., & Heleno, R. H. (2013). Seed dispersal networks in an urban novel ecosystem. European journal of forest research, 132, 887–897. https://doi.org/10.1007/s10342-013-0722-1
Dinno, A. (2017). dunn.test: Dunn’s Test of Multiple Comparisons Using Rank Sums. R package version 1.3.5, https://CRAN.R-project.org/package=dunn.test
Flores-Morales, I. (2023). Diversidad vegetal y animal de los pedregales remanentes de la Zona de Institutos de Investigaciones en Humanidades de Ciudad Universitaria, Ciudad de México, México (Bachelor’s Thesis). Facultad de Ciencias, UNAM. Mexico City.
Glebskiy, Y. (2019). Efecto del conejo castellano (Sylvilagus floridanus) sobre la comunidad vegetal del Pedregal de San Ángel. (M. Sc. Thesis). Facultad de Ciencias, UNAM. México City.
Harrison, R. L. (2012). Ringtail (Bassariscus astutus) ecology and behavior in central New Mexico, USA. Western North American Naturalist, 72, 495–506. https://doi.org/
10.3398/064.072.0407
Kühn, I., Brandl, R., & Klotz, S. (2004). The flora of German cities is naturally species rich. Evolutionary Ecology Research, 6, 749–764.
Mendoza-Hernández, P. E., & Cano-Santana, Z. (2009). Elementos para la restauración ecológica de pedregales: la rehabilitación de áreas verdes de la Facultad de Ciencias en Ciudad Universitaria. In A. Lot, & Z. Cano-Santana (Eds.), Biodiversidad del ecosistema del Pedregal de San Ángel (pp. 523–532). Ciudad de México: Secretaría Ejecutiva de la REPSA-UNAM.
Niu, H., Rehling, F., Chen, Z., Yue, X., Zhao, H., Wang, X. et al. (2023). Regeneration of urban forests as influenced by fragmentation, seed dispersal mode and the legacy effect of reforestation interventions. Landscape and Urban Planning, 233, 104712. https://doi.org/10.1016/j.landurbplan.2023.104712
Niu, H. Y., Xing, J. J., Zhang, H. M., Wang, D., & Wang, X. R. (2018). Roads limit of seed dispersal and seedling recruitment of Quercus chenii in an urban hillside forest. Urban Forestry & Urban Greening, 30, 307–314. https://doi.org/10.1016/j.ufug.2018.01.023
Oksanen, J., Simpson, G., Blanchet, F., Kindt, R., Legendre, P., Minchin, P. et al. (2022). vegan: Community Ecology Package. R package version 2.6-2, https://CRAN.R-project.org/package=vegan
Picazo, G. E. R. C., & García-Collazo, R. (2019). Comparación de la dieta del cacomixtle norteño, Bassariscus astutus de un bosque templado y un matorral xerófilo, del centro de México. Biocyt: Biología, Ciencia y Tecnología, 12, 834–845. https://doi.org/10.22201/fesi.20072082.2019.12.68527
Poglayen-Neuwall, I., & Toweill, D. E. (1988). Bassariscus astutus. Mammalian Species, 327, 1–8.
R Core Team (2022). R: a language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. https://www.R-project.org/
Rodríguez-Estrella, R., Moreno, A. R., & Tam, K. G. (2000). Spring diet of the endemic ring-tailed cat (Bassariscus astutus insulicola) population on an island in the Gulf of California, Mexico. Journal of Arid Environments, 44, 241–246. https://doi.org/10.1006/jare.1999.0579
Rubalcava-Castillo, F. A., Sosa-Ramírez, J., Luna-Ruíz, J. J., Valdivia-Flores, A. G., Díaz-Núñez, V., & Íñiguez-Dávalos, L. I. (2020). Endozoochorous dispersal of forest seeds by carnivorous mammals in Sierra Fría, Aguascalientes, Mexico. Ecology and Evolution, 10, 2991-3003. https://doi.org/10.1002/ece3.6113
Rzedowski, J. (1954). Vegetation of Pedregal de San Ángel. Anales de la Escuela Nacional de Ciencias Biológicas, IPN, Mexico, 8, 59 Endozoochorous 129.
SMN (Servicio Meteorológico Nacional). (2023). Normales climatológicas de la estación 00009071 Colonia Educacion (1991-2020). Consulted on 12 September 2023. https://smn.conagua.gob.mx/tools/RESOURCES/Normales_Climatologicas/Normales9120/df/nor9120_09071.TXT
Swanson, A. C., Conn, A., Swanson, J. J., & Brooks, D. M. (2022). Record of an Urban Ringtail (Bassariscus astutus) Outside of its Typical Geographic Range. Urban Naturalist Notes, 9, 1–6.
Tzintzun-Sánchez, C. L. (2022). Efectos de la urbanización en la distribución geográfica y hábitos alimenticios del cacomixtle norteño (Bassariscus astutus) en la Ciudad de México, México (Bachelor’s Thesis). Facultad de Ciencias, UNAM. Ciudad de México.
Vavrek, M. J. (2011). fossil: palaeoecological and palae-
ogeographical analysis tools. Palaeontologia Electronica, 14:1T. http://palaeo-electronica.org/2011_1/238/index.html
Wania, A., Kühn, I., & Klotz, S. (2006). Plant richness patterns in agricultural and urban landscapes in Central Germany —spatial gradients of species richness. Landscape and Urban planning, 75, 97–110. https://doi.org/10.1016/j.landurbplan.
2004.12.006
Zambrano, L., Rodríguez-Palcios, S., Pérez-Escobedo, M., Gil-Alarcón, G. Camarena, P., & Lot, A. (2016). Reserva Ecológica del Pedregal de San Ángel: atlas de riesgos. Ciudad de Mexico: Secretaría Ejecutiva de la REPSA-UNAM.
Don’t count your eggs before they hatch: differential survival of artificial bird nests in an anthropogenically modified landscape in western Mexico
Dallas R. Levey a, b, Ian MacGregor-Fors c, *
a Stanford University, Department of Biology, 327 Campus Drive, Stanford, California, 94305 USA
b Universidad Nacional Autónoma de México, Instituto de Biología, Tercer Circuito s/n, Ciudad Universitaria, Coyoacán, 04510 Mexico City, Mexico
c University of Helsinki, Faculty of Biological and Environmental Sciences, Ecosystems and Environment Research Programme, Niemenkatu 73, FI-15140, Lahti, Finland
*Corresponding author: ian.macgregor@helsinki.fi (I. MacGregor-Fors)
Received: 19 October 2023; accepted: 1 February 2024
Abstract
Native habitat conversion to urban and agricultural areas represents conservation concerns for habitat quality and the breeding success of birds. In tropical areas facing regular deforestation of at-risk habitats, changes may occur to bird and nest predator communities that influence contradictory trends in breeding success. To assess the value of working lands for birds, we placed 100 artificial nests in 5 habitat types of varying human footprint, including a tropical dry forest reserve, a biological research station, croplands, and 2 urban towns. We report a clear decline in survival from the forest to urban towns. Habitat type explained the variation in nest survival probabilities over nest height, elevation, or time of nest exposure. Reducing the structural and compositional contrast of habitat and landscape vegetation between tropical dry forest and working lands represent valuable conservation actions for increasing habitat quality for birds.
Keywords: Croplands; Bird nest predation; Habitat quality; Jalisco; Plasticine eggs; Tropical dry forest; Urbanization
© 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/).
No cuentes los huevos antes de que eclosionen: supervivencia diferencial de nidos artificiales de aves en un paisaje antropogénicamente modificado en el oeste de México
Resumen
La conversión de hábitats nativos en áreas urbanas y campos agrícolas representa problemas de conservación para la calidad del hábitat y el éxito reproductivo de las aves. La deforestación constante de hábitats en riesgo puede cambiar las comunidades de aves y los depredadores de los nidos, lo que puede influir en su éxito reproductivo. Para evaluar el valor de los hábitats dentro de un paisaje antropogénicamente modificado en el éxito reproductivo de las aves, colocamos 100 nidos artificiales en 5 hábitats con diferentes niveles de actividades humanas, incluyendo una reserva de bosque tropical caducifolio, una estación de investigación biológica, campos agrícolas y 2 pueblos urbanos. Encontramos una clara disminución en la supervivencia de nidos artificiales desde el bosque tropical caducifolio hasta los pueblos urbanos. El tipo de hábitat fue la variable que mejor explicó la variación en las probabilidades de supervivencia de los nidos artificiales en comparación con la altura del nido, la elevación y el tiempo de exposición del nido. Reducir el contraste dentro del paisaje en la estructura de la vegetación entre la reserva y los hábitats dentro del paisaje modificado representan acciones de conservación importantes para aumentar la calidad del hábitat para las aves.
Palabras clave: Campos agrícolas; Depredación de nidos de aves; Calidad de hábitat; Jalisco; Huevos de plastilina; Selva seca tropical; Urbanización
Introduction
Urbanization and the conversion of native habitat to agricultural land represent key factors in the long-term conservation of bird biodiversity (Aronson et al., 2014; Kehoe et al., 2017). In the tropics, urbanization and agriculture have led to the degradation and destruction of native vegetation and the reconfiguration of landscapes, causing stark contrast in habitat complexity between remnant vegetation and agricultural and urban areas that pose risks to bird biodiversity (Filloy et al., 2019; Fischer et al., 2015; Maas et al., 2016). Biodiverse tropical regions suffer some of the highest rates of urbanization and native habitat transformation (Estrada et al., 2020), which represent pressing challenges for the conservation of bird populations.
A key component of bird biodiversity and population monitoring in tropical landscapes with high rates of natural habitat transformation includes the evaluation of breeding ecology (DeGregorio et al., 2016). Reduced vegetative complexity and the exchange of native plants with non-native plants, both common attributes of agricultural and urban areas (Chace & Walsh, 2006), tend to negatively impact bird species with highly sensitive breeding requirements tied to native vegetation (Maas et al., 2016). Vegetation change in the tropics alters biotic —e.g., nest predation pressure and reduction of nest locations— and abiotic conditions —e.g., increased nest exposure, higher temperatures, and brighter conditions—, leading to potential direct and indirect influences on avian breeding ecology in species that depend on native plants and vegetation structure for nesting (Estrada et al., 2002; Rivera-López & MacGregor-Fors, 2016; Tellería & Díaz, 1995; Zuñiga-Palacios et al., 2021). Meanwhile, certain bird species may be positively impacted by or able to acclimate to novel conditions (DeGregorio et al., 2016; Kurucz et al., 2021; Latif et al., 2012), underlining the semi-permeable ecological filter that is applied to nesting birds in human-modified tropical landscapes and the importance of evaluating bird breeding ecology in different types of transformed land (MacGregor-Fors, 2010; MacGregor-Fors et al., 2022).
Nest predation represents a powerful force on bird breeding success and population dynamics (DeGregorio et al., 2016). A consequence of native habitat conversion to more urban or agricultural areas includes changes to nest predator communities, and the ability of bird species to adapt to these changes will ultimately determine whether disturbed areas offer viable habitat for native levels of biodiversity (DeGregorio et al., 2016; Latif et al., 2012). Urban and agricultural areas tend to have lower vegetation cover as forested habitats, leading to different natural predator abundance and nest visibility to predators, representing powerful determinants of breeding success for bird species that use transformed land (López-Flores et al., 2009; Martin, 1993; Zuñiga-Palacios et al., 2021). More disturbed areas may lead to a reduction of nest predation pressure due to the absence of native nest predators that have a low tolerance for human activity (Kurucz et al., 2021; Pretelli et al., 2023). A possible caveat to lower predation pressure from typical native predators in urban areas include birds that are habitat and foraging generalists (Estrada et al., 2002; Martin, 1995; Rivera-López & MacGregor-Fors, 2016), mammals that are attracted to anthropogenic food sources (Fischer et al., 2012), and increased access to nests by people that manipulate and destroy nests and eggs (López-Flores et al., 2009).
To assess the survivorship of bird nests, we placed artificial bird nests in 5 habitat types with increasing degrees of human disturbance and habitat modification, including 1) a tropical dry forest reserve (TDF hereafter), 2) the Chamela Biological Research Station (CBRS) grounds, embedded in the Chamela-Cuixmala Biosphere Reserve, 3) croplands (CL), 4) Careyes (CAR), a small and heavily built-up town, and 5) Emiliano-Zapata (ZAP), a larger town. While a recent meta-analysis questions the efficacy of artificial nest studies in determining nest survival probabilities in urban areas relative to natural nests (Vincze et al., 2017), the feasibility of finding sufficient numbers of natural nests in heavily built-up urban areas (i.e., outside of urban parks and green spaces) makes the use of artificial nests necessary. We controlled for important variables that may influence predation rates, such as nest size and height, to focus on habitat-level variations in nest survivorship and the impacts of urban and agricultural areas in the working landscape. Such landscapes are common in the tropical areas of Mexico (Levey et al., 2023), where existing reserves are surrounded by a working landscape with non-native vegetation that contrasts highly with native areas (Levey & MacGregor-Fors, 2021; Levey et al., 2021; MacGregor-Fors & Schondube, 2011; Vázquez-Reyes et al., 2017). Efforts to evaluate the impacts on breeding ecology in these working tropical landscapes are needed to supplement a thin body of work (Estrada et al., 2002; López-Flores et al., 2009; Zuñiga-Palacios et al., 2021) and determine the risks that urban and agricultural areas present for breeding birds. We expected nest survivorship to be lower in CL, CAR, and ZAP relative to the conserved TDF reserve and the CBRS due to higher exposure of nests and greater visibility for predators due to reduced vegetation complexity and density (Estrada et al., 2002; López-Flores et al., 2009; Zuñiga-Palacios et al., 2021).
Materials and methods
We conducted our study in a landscape between the Chamela-Cuixmala Biosphere Reserve (19°29’57.5” N, 105°02’41.6” W) and the town Emiliano Zapata (19°23’16.6” N, 104°57’50.1” W) in the Municipality La Huerta (population: 23,258; INEGI, 2020) on the Pacific coast of Jalisco, Mexico (Fig. 1). Historically, native vegetation cover in the region consisted primarily of tropical dry forest, which consists of deciduous forest with a mean canopy height of 12 m, a dense understory (Rzedowski, 2006), and strong phenological changes due to highly seasonal rainfall in the region (Durán et al., 2002). Other forest types exist in areas with more regular water availability, including semi-deciduous (mean canopy height of 20 m) and mangrove forests (Durán et al., 2002). After a period of increased human occupation and agricultural expansion from 1950-1970, large cover of tropical dry forest and other native forest types in lower elevation zones were converted to small towns and agricultural lands, linked by paved and unpaved roads, creating a landscape mosaic of native and non-native vegetation types (Maass et al., 2005).
In this landscape, we selected 5 habitat types with varying degrees of urban and agricultural disturbance for artificial nest placement: 1) TDF, with closed canopy cover and dense understory, 2) CBRS, which consists of moderately built-up 1.4 ha area embedded within the Chamela-Cuixmala Biosphere Reserve, 3) CL, consisting of fields of small, herbaceous plants such as maize (Zea mays), squash (Cucurbita spp.), chili pepper (Capsicum spp.), watermelon (Citrullus lanatus), and beans (Phaseolus spp.) located in the southern edge of the study area (Maass et al., 2005), 4) CAR (19°26’36.15” N, 105°1’49.65” W), a small town with heavy built-up cover, and 5) ZAP, a large town with less built-up cover than CAR (Fig. 1). Both elevations (MSL) of the TDF and CBRS sampling areas were slightly higher than the other habitat types. The TDF and CBRS sampling areas are also in closer proximity to each other than the other sampling locations. We included both habitat categories due to the higher human presence at the Biological Station, the noise generated by people and activities at the station, and a higher density of paved roads that could influence the occupancy of bird and mammals that respond positively to increasing human footprint (Rivera-López & MacGregor-Fors, 2016). Potential bird nest predators in the study region included White-throated Magpie-jays and San Blas jays (Calocitta formosa and Cyanocorax sanblasianus), Great-tailed grackles (Quiscalus mexicanus), mammals (e.g., Nasua narica, domestic dogs and cats, rodents, possums, and Procyon lotor), and diverse reptiles.
We used a mixture of plant fibers, twigs, and mud from the nest location to create bird nests in the shape of open plant fiber nest cups large enough to hold both the clay and quail eggs. We created open cup nests since many species in the study area construct nests in similar ways (e.g., Cyanocorax sanblasianus, Peucaea ruficauda, and Turdus rufopalliatus; Mendoza-Rodríguez et al., 2010) and due to the ease of creating such a nest shape. We placed each nest ~ 2 m above ground to control the tendency of nest height placement to affect predation rates (DeGregorio et al., 2016). We placed 1 Japanese Quail (Coturnix japonica) commercial egg and 1 clay egg of similar size for a total of 2 eggs in each nest (Bayne et al., 1997; Estrada et al., 2002). We used clay since it is a malleable material that preserves markings from predation events and has negligible influence on predation rates (Bayne & Hobson, 1999; Bayne et al., 1997). We used commercial quail eggs due to their small size and color speckling that best mimicked natural terrestrial bird eggs relative to domestic chicken eggs and their availability in the study region. We used both a real and clay egg to provide stimulus for a wider range of predators than clay eggs alone and to capture predation event evidence if we could not perceive markings on the quail egg from smaller nest predators (Bayne et al., 1997; Estrada et al., 2002). We used rubber gloves to prevent leaving a human scent when handling nest materials and eggs (Estrada et al., 2002).
Figure 1. Region of study in the state of Jalisco in western Mexico. We placed artificial bird nests at the localities marked with a black dot and text, including ‘Forests’ (tropical dry forest of the Chamela-Cuixmala Biosphere Reserve), ‘Biology station’ (the Chamela Biological Research Station), ‘Cropfields’ (herbaceous crop plots), ‘Careyes’ (a small, heavily built-up town), and ‘Zapata’ (a large, less built-up town). Nest locations within the marked localities by at least 250 m to increase spatial independence.
We placed 20 artificial nests in each of the 5 habitat types for a total of 100 nests. Nests were exposed for a total of 12 days (April 30 – May 11, 2009), and we checked nests at 3-day intervals for a total of 4 nest visits. We considered nests as failed if the eggs were missing or if there were indications of a predation event on either the clay or quail egg, including scratches, bite marks, or perforations. We removed nests with signs of predation from the sample locations. We considered nests successful if there were no markings on either the clay or quail egg.
Figure 2. Survival probability with 95% confidence intervals from the Known Fate analysis in MARK of the artificial nests in the conserved tropical dry forest (TDF in the figure), Chamela Biological Research Station (CBRS), cropland (CL), the town of Careyes (CAR), and the town of Zapata (ZAP).
We used the program MARK (White & Burnham, 1999) to perform a known fate analysis using our nest check interval to calculate the probability of survivorship of each nest (Dinsmore & Dinsmore, 2007), using the covariables nest height (m), elevation (m asl), habitat type, and time of nest exposure to generate the models. We included the nest height variable in analyses despite controlling the height at 2-m to check for potential interactions with other covariates. We included elevation in our models to account for slight elevation differences between nest site locations and the tendency of lower elevation areas to have higher cover of agricultural and urban areas (Maass et al., 2005). We included habitat type to determine the differences between certain habitat types on artificial nest survival. Finally, we included the time of nest exposure since the likelihood of nest survival is tied to the amount of time eggs are exposed to predators (Dinsmore & Dinsmore, 2007). We ranked the models by parsimony using the adjusted Akaike’s Information Criterion for small sample sizes (AICc; Hurvich & Tsai, 1989). We selected the models that best fit our data by calculating the differences in AICc values (ΔAICc) and choosing those with ΔAICc values less than 2 units from the most parsimonious model (Burnham & Anderson, 2002).
Results
We recorded 82 preyed upon nests of 100 total, including 37 (45.1%) instances of bird predation, 30 (36.6%) instances of unknown predation, 10 (12.2%) instances of egg removal or manipulation by humans, 3 (3.7%) instances of rodent predation, and 2 (2.4%) instances of reptile predation. We recorded 18 nests with no predation signs, with the majority remaining in TDF (44.4%), followed by CBRS (27.8%), CL (22.2%), and ZAP (5.6%). No nests placed in CAR survived the observation period. Nest survival probability was 0.38 (95% CI: 0.26-0.48) in TDF, 0.27 (95% CI: 0.17-0.33) in CBRS, 0.25 (95% CI: 0.15-0.35) in CL, 0.06 (95% CI: 0.01-0.13) in ZAP, and 0.0 in CAR (Fig. 2). The most parsimonious model to explain the variation in nest survival probabilities included the lone covariable habitat, followed closely by the combination of habitat and height (Table 1).
Table 1
Model output from the Known Fate analysis in MARK. Covariates used in the models include habitat, nest height (controlled at 2 m above ground level), elevation, and time of nest exposure.
Model | AICc | ΔAICc | AICc weight | Model likelihood | Parameters | Deviance |
Habitat | 262.62 | 0.00 | 0.44 | 1.00 | 5 | 252.32 |
Habitat + height | 263.96 | 1.35 | 0.22 | 0.51 | 6 | 251.55 |
Habitat + elevation | 264.49 | 1.87 | 0.17 | 0.39 | 6 | 252.08 |
Habitat + height + elevation | 265.90 | 3.29 | 0.09 | 0.19 | 7 | 251.35 |
Elevation | 267.21 | 4.59 | 0.04 | 0.10 | 5 | 256.91 |
Height + elevation | 268.74 | 6.12 | 0.02 | 0.05 | 6 | 256.32 |
Time of nest exposure | 270.31 | 7.69 | 0.01 | 0.02 | 4 | 262.11 |
Height | 272.26 | 9.64 | 0 | 0.01 | 5 | 261.96 |
Discussion
The impacts of bird nest predation along habitat disturbance gradients vary depending on the severity of habitat modification and the biotic and abiotic conditions of transformed land (Vincze et al., 2017). Novel biotic and abiotic conditions in urban and agricultural settings heavily contrast with native habitat, representing important influences on bird breeding success and, ultimately, biodiversity conservation (DeGregorio et al., 2016). We report a clear decline in the survival probabilities of artificial bird nests throughout a gradient of urban intensity between a conserved tropical dry forest and the largest town.
TDF, the most conserved habitat in the disturbance gradient, had the highest artificial nest survival probability among all studied habitats. This finding is consistent with other artificial nest studies from the tropical Americas that show greater vegetation cover offers increased survival odds by concealing nests more effectively from predators, both within forests with seasonal leaf cover (Vega-Rivera et al., 2009) and relative to more open areas (Estrada et al., 2002; López-Flores et al., 2009). TDF contains a dense understory of vegetation and a closed canopy with darker lighting, which may be a key factor in the detection of nests by predators (Estrada et al., 2002; Vázquez et al., 2021). While some studies have found that conserved areas either have similar or lower nest survival probabilities than in urban settings due to changes in predator abundance and composition (DeGregorio et al., 2016; Fischer et al., 2012; Zuñiga-Palacios et al., 2021), local factors in this heterogeneous landscape with various habitat types likely favor ample distribution of potential nest predators (e.g., urban birds, domesticated cats, and dogs) in urban areas (Estrada et al., 2002; López-Flores et al., 2009; Rivera-López & MacGregor-Fors, 2016).
Outside of the conserved TDF habitat, CL showed near-equal nest survival probabilities as the CBRS, which were lower than in TDF. Our results indicate that even small (< 2 ha), moderately built-up areas embedded in conserved habitat may increase the likelihood of nest predation to levels found in agricultural land. Synonymous with development is the opening of forest habitat, leading to new abiotic conditions and biotic stimulus that may influence breeding success in birds (Patten & Smith-Patten, 2012; Shochat et al., 2010). In our study area, CBRS has attracted several bird species that are opportunistic omnivores and often associated with open habitats, such as the Great-tailed Grackle (Quiscalus mexicanus; MacGregor-Fors et al., 2009). Also attracted to this habitat and CL are potential nest predators such as the White-nosed Coati (Nasua narica) and Common Raccoon (Procyon lotor), which have been documented to predate bird nests (Estrada et al., 2002; Menezes & Marini, 2017; Robinson et al., 2005). Snakes, which occur at similar compositions inside and outside the reserve, may exhibit increased activity at edge habitats (Chalfoun et al., 2002; Suazo-Ortuño et al., 2008; Vetter et al., 2013). These changes to the nest predator communities in CBRS and CL could have important implications on bird breeding success (DeGregorio et al., 2016), and continued urbanization of these areas may continue to decrease nest survival probabilities to the levels of heavily built-up towns.
The built-up areas along the urbanization gradient in our study had significantly lower nest survival probabilities than the other studied habitats. Urbanization and loss of native vegetation have been shown to negatively influence the survival of bird nests in previous studies (Rivera-López & MacGregor-Fors, 2016; Thorington & Bowman, 2003), and a potential mechanism includes the introduction of novel predation pressures, such as domesticated cats (Patterson et al., 2016), dogs (Zuñiga-Palacios et al., 2021) and humans (López-Flores et al., 2009). While it has been shown that urban areas may increase nest survival and breeding success in birds (Fischer et al., 2012; Kurucz et al., 2021), the urban areas in our study area presented an overwhelming amount of novel predation pressures that are not present in the other studied habitats (Chace & Walsh, 2006; López-Flores et al., 2009), highlighting the importance of evaluating changes in the communities of bird nest predators along habitat disturbance gradients (DeGregorio et al., 2016). Conserving and restoring degraded areas within working landscapes and urban centers through measures such as live fencing, remnant forest preservation, and educational programs on bird breeding ecology may provide vital nesting habitat and increase bird breeding success (Bocz et al., 2017; Zuñiga-Palacios et al., 2021).
Acknowledgements
We thank the Estación de Biología Chamela (Instituto de Biología, UNAM) for granting permission to place artificial nests in the biosphere reserve. We thank Carlos Lara for editing and three anonymous reviewers for comments that enhanced the quality and clarity of the manuscript. We thank Michelle García-Arroyo for creating the study area map. DRL received a Master’s scholarship from Conahcyt (grant number 964233) as part of the Posgrado en Ciencias Biológicas of the Universidad Nacional Autónoma de México.
References
Aronson, M. F. J., La Sorte, F. A., Nilon, C. H., Katti, M., Goddard, M. A., Lepczyk, C. A. et al. (2014). A global analysis of the impacts of urbanization on bird and plant diversity reveals key anthropogenic drivers. Proceedings of the Royal Society B: Biological Sciences, 281, 20133330. https://doi.org/10.1098/rspb.2013.3330
Bayne, E. M., & Hobson, K. A. (1999). Do clay eggs attract predators to artificial nests? Journal of Field Ornithology, 70, 1–7.
Bayne, E. M., Hobson, K. A., & Fargey, P. (1997). Predation on artificial nests in relation to forest type: contrasting the use of quail and plasticine eggs. Ecography, 20, 233–239. https://doi.org/10.1111/j.1600-0587.1997.tb00366.x
Bocz, R., Szép, D., Witz, D., Ronczyk, L., Kurucz, K., & Purger, J. J. (2017). Human disturbances and predation on artificial ground nests across an urban gradient. Animal Biodiversity Conservation, 40, 153–157. https://doi.org/10.32800/abc.2017.40.0153
Burnham, K. P., & Anderson, D. R. (2002). Model selection and multimodel inference. New York: Springer-Verlag.
Chace, J. F., & Walsh, J. J. (2006). Urban effects on native avifauna: a review. Landscape and Urban Planning, 74, 46–69. https://doi.org/10.1016/j.landurbplan.2004.08.007
Chalfoun, A. D., Thompson, F. R., & Ratnaswamy, M. J. (2002). Nest predators and fragmentation: a review and meta-analysis. Conservation Biology, 16, 306–318.
DeGregorio, B. A., Chiavacci, S. J., Benson, T. J., Sperry, J. H., & Weatherhead, P. J. (2016). Nest predators of North American birds: continental patterns and implications. Bioscience, 66, 655–665. https://doi.org/10.1093/biosci/biw071
Dinsmore, S. J., & Dinsmore, J. J., 2007. Modeling avian nest survival in program MARK. Studies in Avian Biology, 34, 73–83.
Durán, E., Balvanera, P., Lott, E., Segura, G., Pérez-Jiménez, A., Islas, Á. et al. (2002). Estructura, composición y dinámica de la vegetación. In F. A. Noguera, J. H. Vega-Rivera, A. N. García-Aldrete, & M. Quesada-Avendaño (Eds.), Historia natural de Chamela (pp. 443–472). México D.F.: Instituto de Biología, Universidad Nacional Autónoma de México.
Estrada, A., Garber, P. A., & Chaudhary, A. (2020). Current and future trends in socio-economic, demographic and governance factors affecting global primate conservation. PeerJ, 8, e9816. https://doi.org/10.7717/peerj.9816
Estrada, A., Rivera, A., & Coates-Estrada, R. (2002). Predation of artificial nests in a fragmented landscape in the tropical region of Los Tuxtlas, Mexico. Biological Conservation, 106, 199–209. https://doi.org/10.1016/S0006-3207(01)00246-4
Filloy, J., Zurita, G. A., & Bellocq, M. I. (2019). Bird diversity in urban ecosystems: the role of the biome and land use along urbanization gradients. Ecosystems, 22, 213–227. https://doi.org/10.1007/s10021-018-0264-y
Fischer, J. D., Cleeton, S. H., Lyons, T. P., & Miller, J. R. (2012). Urbanization and the predation paradox: the role of trophic dynamics in structuring vertebrate communities. Bioscience, 62, 809–818. https://doi.org/10.1525/bio.2012.62.9.6
Fischer, J. D., Schneider, S. C., Ahlers, A. A., & Miller, J. R. (2015). Categorizing wildlife responses to urbanization and conservation implications of terminology. Conservation Biology, 29, 1246–1248. https://doi.org/10.1111/cobi.12451
Hurvich, C. M., & Tsai, C. L. (1989). Regression and time series model selection in small samples. Biometrika, 76, 297–307. https://doi.org/10.1093/biomet/76.2.297
INEGI (Información Estadística y Geográfica). (2020). Censo de población y vivienda 2020. Recovered on 12 May 2021 from: https://www.inegi.org.mx/programas/ccpv/2020/
Kehoe, L., Romero-Muñoz, A., Polaina, E., Estes, L., Kreft, H., & Kuemmerle, T. (2017). Biodiversity at risk under future cropland expansion and intensification. Nature Ecology and Evolution, 1, 1129–1135. https://doi.org/10.1038/s41559-017-0234-3
Kurucz, K., Purger, J. J., & Batáry, P. (2021). Urbanization shapes bird communities and nest survival, but not their food quantity. Global Ecology and Evolution, 26, e01475. https://doi.org/10.1016/j.gecco.2021.e01475
Latif, Q. S., Heath, S. K., & Rotenberry, J. T. (2012). How avian nest site selection responds to predation risk: testing an ‘adaptive peak hypothesis’. Journal of Animal Ecology, 81, 127–138. https://doi.org/10.1111/j.1365-2656.2011.01895.x
Levey, D. R., Estrada, A., Enríquez, P. L., & Navarro-Sigüenza, A. G. (2021). The importance of forest-nonforest transition zones for avian conservation in a vegetation disturbance gradient in the northern Neotropics. Tropical Conservation Science, 14, 1–14. https://doi.org/10.1177/19400829211008087
Levey, D. R., & MacGregor-Fors, I. (2021). Neotropical bird communities in a human-modified landscape recently affected by two major hurricanes. Avian Conservation and Ecology, 16, art9. https://doi.org/10.5751/ACE-01920-160209
Levey, D. R., Patten, M. A., & Estrada, A. (2023). Bird species occupancy trends in southeast Mexico over 1900–2020: accounting for sighting record absences. Journal of Animal Ecology, 92, 606–618. https://doi.org/10.1111/1365-2656.13871
López-Flores, V., MacGregor-Fors, I., & Schondube, J. E. (2009). Artificial nest predation along a Neotropical urban gradient. Landscape and Urban Planning, 92, 90–95. https://doi.org/10.1016/j.landurbplan.2009.03.001
Maas, B., Karp, D. S., Bumrungsri, S., Darras, K., Gonthier, D., Huang, J. C. C. et al. (2016). Bird and bat predation services in tropical forests and agroforestry landscapes. Biological Reviews, 91, 1081–1101. http://doi.wiley.com/10.1111/brv.12211
Maass, J. M., Balvanera, P., Castillo, A., Daily, G. C., Mooney, H. A., Ehrlich, P. et al. (2005). Ecosystem services of tropical dry forests: insights from long-term ecological and social research on the Pacific coast of Mexico. Ecology and Society, 10, art17. https://doi.org/10.5751/es-01219-100117
MacGregor-Fors, I. (2010). How to measure the urban-wildland ecotone: redefining ‘peri-urban’ areas. Ecological Research, 25, 883–887. https://doi.org/10.1007/s11284-010-0717-z
MacGregor-Fors, I., García-Arroyo, M., & Quesada, J. (2022). Keys to the city: an integrative conceptual framework on avian urban filtering. Journal of Urban Ecology, 8, juac026. https://doi.org/10.1093/jue/juac026
MacGregor-Fors, I., & Schondube, J. E. (2011). Use of tropical dry forests and agricultural areas by Neotropical bird communities. Biotropica, 43, 365–370. http://doi.wiley.com/10.1111/j.1744-7429.2010.00709.x
MacGregor-Fors, I., Vázquez, L., Vega-Rivera, J. H., & Schondube, J. E. (2009). Non-exotic invasion of Great-tailed Grackles (Quiscalus mexicanus) in a tropical dry forest reserve. Aredea, 97, 367–369. https://doi.org/10.5253/078.097.0312
Martin, T. E. (1993). Nest predation among vegetation layers and habitat types: revising the dogmas. American Naturalist, 141, 897–913. https://doi.org/10.1086/285515
Martin, T. E. (1995). Avian life history evolution in relation to nest sites, nest predation, and food. Ecological Monographs, 65, 101–127. https://doi.org/10.2307/2937160
Mendoza-Rodríguez, V., Vega-Rivera, J. H., Medina-Montaño, I., & Campos-Cerda, F. (2010). Response of birds in tropical deciduous forest to Brown-headed Cowbirds (Molothrus ater). The Southwestern Naturalist, 55, 390–393. https://doi.org/10.1894/MH-46.1
Menezes, J. C. T., & Marini, M. Â. (2017). Predators of bird nests in the Neotropics: a review. Journal of Field Ornithology, 88, 99–114. https://doi.org/10.1111/jofo.12203
Patterson, L., Kalle, R., & Downs, C. (2016). Predation of artificial bird nests in suburban gardens of KwaZulu-Natal, South Africa. Urban Ecosystems, 19, 615–630. http://dx.doi.org/10.1007/s11252-016-0526-4
Pretelli, M. G., Cavalli, M., Chiaradia, N. M., Cardoni, A., & Isacch, J. P. (2023). Location matters: survival of artificial nests is higher in small grassland patches and near the patch edge. Ibis, 165, 111–124. https://doi.org/10.1111/ibi.13128
Rivera-López, A., & MacGregor-Fors, I. (2016). Urban predation:
a case study assessing artificial nest survival in a Neotro-
pical city. Urban Ecosystems, 19, 649–655. http://dx.doi.org/10.1007/s11252-015-0523-z
Robinson, W. D., Styrsky, J. N., & Brawn, J. D. (2005). Are artificial bird nests effective surrogates for estimating predation on real bird nests? A test with tropical birds. Auk, 122, 843–852. https://doi.org/10.1093/auk/122.3.843
Rzedowski, J. (2006). Vegetación de México. Edicion digital. Comisión Nacional para el Conocimiento y Uso de la Biodiversidad. México, Ciudad de México. https://www.biodiversidad.gob.mx/publicaciones/librosDig/pdf/VegetacionMxPort.pdf
Suazo-Ortuño, I., Alvarado-Díaz, M., & Martínez-Ramos, M. (2008). Effects of conversion of dry tropical forest to agricultural mosaic on herpetofaunal assemblages. Conservation Biology, 22, 362-374. https://doi.org/10.1111/
j.1523-1739.2008.00883.x
Shochat, E., Lerman, S. B., Anderies, J. M., Warren, P. S., Faeth, S. H., & Nilon, C. H. (2010). Invasion, competition, and biodiversity loss in urban ecosystems. Bioscience, 60, 199–208. https://doi.org/10.1525/bio.2010.60.3.6
Tellería, J. L., & Díaz, M. (1995). Avian nest predation in a large natural gap of the Amazonian rainforest. Journal of Field Ornithology, 66, 343–351.
Thorington, K. K., & Bowman, R. (2003). Predation rate on artificial nests increases with human housing density in suburban habitats. Ecography, 26, 188–196. https://doi.org/10.1034/j.1600-0587.2003.03351.x
Vázquez-Reyes, L. D., Arizmendi, M. C., Godínez-Álvarez, O. H., & Navarro-Sigüenza, A. G. (2017). Directional effects of biotic homogenization of bird communities in Mexican seasonal forests. Ornithological Applications, 119, 275–288. https://doi.org/10.1650/CONDOR-16-116.1
Vazquez, M. S., Zamora-Nasca, L. B., Rodríguez-Cabal, M. A., & Amico, G. C., 2021. Interactive effects of habitat attributes and predator identity explain avian nest predation patterns. Emu – Austral Ornithology, 121, 250–260. https://doi.org/10.1080/01584197.2021.1928519
Vega-Rivera, J. H., Medina-Montaño, I., Rappole, J. & Campos-Cerda, F. (2009). Testing the importance of nest concealment: does timing matter? Journal of Field Ornithology, 80, 303–307. https://doi.org/10.1111/j.1557-9263.2009.00234.x
Vetter, D., Rücker, G., & Storch, I. (2013). A meta-analysis of tropical forest edge effects on bird nest predation risk: edge effects in avian nest predation. Biological Conservation, 159, 382–395. https://doi.org/10.1016/j.biocon.2012.12.023
Vincze, E., Seress, G., Lagisz, M., Nakagawa, S., Dingemanse, N. J., & Sprau, P. (2017). Does urbanization affect predation of bird nests? a meta-analysis. Frontiers in Ecology and Evolution, 5, 29. https://doi.org/10.3389/fevo.2017.00029
White, G. C., & Burnham, K. P. (1999). Program MARK: survival estimation from populations of marked animals. Bird Study, 46 (Suppl.), S120–S139. https://doi.org/10.1080/00063659909477239
Zuñiga-Palacios, J., Corcuera, P., & Almazán-Núñez, R. C. (2021). Living fences decrease the edge effect on nest predation in a tropical dry forest landscape: evidence from an experiment using artificial nests. Agroforestry Systems, 95, 547–558. https://doi.org/10.1007/s10457-021-00603-z
Ephemeral and intermittent xeroriparian systems arekeystone habitats for bird communities during thenon-breeding season in a Mexican semiarid landscape
Mónica E. Riojas-López a, Eric Mellink b, *, Moisés Montes-Olivares b
a Universidad de Guadalajara, Centro Universitario de Ciencias Biológicas y Agropecuarias, Departamento de Ecología, C. Ramón Padilla Sánchez Núm. 2100, 45100 Zapopan, Jalisco, Mexico
b Centro de Investigación Científica y de Educación Superior de Ensenada, Departamento de Biología de la Conservación, Carretera Ensenada-Tijuana Núm. 3918, Zona Playitas, 22860 Ensenada, Baja California, Mexico
*Corresponding author: emellink@cicese.mx (E. Mellink)
Received: 23 October 2023; accepted: 9 April 2024
Abstract
Intermittent and ephemeral xeroriparian systems cover less than 1% of continental North America and are critical for wildlife in arid and semi-arid areas but are understudied and absent from conservation plans. We report the diversity of birds in 3 xeroriparian systems of the Mexican Altiplano during the non-breeding season and the habitat variables that influence them. Of the 48 documented species in this study, we have recorded 15 only in these systems, throughout our long-time research in the region. Bird communities were positively influenced by minimum and maximum height of shrubs and trees and negatively by canopy cover. The communities were grouped in one gradient from lower richness in rocky, entrenched streams, with closed canopy and little herbaceous vegetation, to greater richness in wide, open streams, with abundant herbaceous plants, and in a second gradient, from insectivorous to granivorous birds. Our study covered habitats not considered in other similar studies in Mexico and revealed that at the landscape level, ephemeral and intermittent xeroriparian systems could play a crucial role in conservation given that the systems studied covered approximately 0.1% of the area but hosted 20% of the region’s land bird species and, among migrants, especially Spring migrants.
Keywords: Arid lands; Semiarid lands; Llanos de Ojuelos; Anthropized landscapes; Migratory birds
© 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/).
Los sistemas xeroribereños efímeros e intermitentes son hábitats clave para comunidades de aves en la temporada no reproductiva en un paisaje semiárido mexicano
Resumen
Los hábitats xeroribereños intermitentes y efímeros cubren menos de 1% de la superficie continental de Norteamérica y son críticos para la fauna silvestre de zonas áridas y semiáridas, pero están poco estudiados y ausentes de planes de conservación. Reportamos la diversidad de aves en 3 sistemas xeroribereños del Altiplano Mexicano durante la temporada no reproductiva y las variables del hábitat que influyen. De 48 especies documentadas, hemos registrado 15 solo en sistemas xiroribereños en muchos años de investigación en la región. Arbustos y árboles más altos tuvieron influencia positiva en la comunidad de aves, mientras que doseles cerrados la tuvieron negativamente. Las comunidades se agruparon de menor riqueza en arroyos rocosos y encañonados con dosel cerrado y poca vegetación herbácea, a mayor riqueza en arroyos amplios y abiertos con abundantes herbáceas, y en un segundo gradiente, de aves insectívoras a granívoras. Nuestro estudio cubrió hábitats no considerados en otros trabajos similares en México y reveló que a nivel de paisaje, los sistemas xeroribereños efímeros e intermitentes podrían ser importantes en la conservación: los sistemas estudiados cubrían aproximadamente 0.1% del área, pero albergaron 20% de las especies de aves terrestres de la región, y entre especies migrantes, especialmente las de primavera.
Palabras clave: Zonas áridas; Zonas semiáridas; Llanos de Ojuelos; Paisajes antropizados; Aves migratorias
Introduction
Riparian systems are plant communities that develop as a result of perennial, intermittent, or ephemeral surface or subsurface water (Krueper, 1993). These systems are one of the rarest habitats in North America and cover less than 5% of the continental land mass (Krueper, 2000). Despite their rarity, throughout the world, riparian systems are extremely important because of their disproportionate contribution, relative to area, for biodiversity conservation (Arizmendi et al., 2008; Carlisle et al., 2009; Hinojosa-Huerta et al., 2013; Kirkpatrick et al., 2009; Knopf, 1985; Krueper 1993, 1996, 2000; Seymour & Simmons, 2008; Skagen et al., 1998; Wilson in Knopf et al., 1988). They also contribute to enhance connectivity in fragmented landscapes particularly for resident and non-migrating birds (Şekercioǧlu et al., 2015).
Dryland riparian systems are known as xeroriparian, and whether the streams that originate them are perennial, or non-perennial, they are notoriously different from the surrounding landscape. In the western United States, in the 1980s they covered < 1% of the land (Knopf et al., 1988). However, they are very important for wildlife in semiarid and arid regions, and support much of the biotic diversity in semiarid and arid southwestern USA (Sánchez-Montoya et al., 2017; Szaro & Jakle, 1985), sometimes having bird population densities and species diversity as much as 5 to 10 times those of nearby desert non-riparian systems (Johnson & Haight, 1985; Levick et al., 2008). In these regions, migrating birds depend on water, habitat, and food that are restricted spatially and temporally (Carlisle et al., 2009). Up to 70% of all bird species use riparian systems in drylands at some point in their life cycle (Krueper, 1996), and > 60% of the neotropical migratory bird species use them either as stopover areas or as breeding habitats (Kirkpatrick et al., 2009; Krueper, 1993; Skagen et al., 1998).
Although non-perennial streams are the most widespread flowing-water ecosystem throughout the world (Datry et al., 2017), ecological studies on xeroriparian systems had focused mostly on permanent streams (Hinojosa-Huerta et al., 2013; Neate-Clegg et al., 2021; Szaro & Jakle, 1985). Overall, riparian systems created and maintained by intermittent and ephemeral streams are understudied and the scientific literature on their ecological role is very limited (Datry et al., 2017; Levick et al., 2008; McDonough et al., 2011; Sánchez-Montoya et al., 2017). Not only are they understudied, but they also are poorly considered in conservation planning. For example, intermittent and ephemeral streams are recognized in California´s “Riparian Bird Conservation Plan” (Riparian Habitat Joint Venture, 2004), but only perennial ones are included in its actions. Such neglect of largely intermittent or ephemeral riparian systems can lead to serious shortcomings in conserving biodiversity.
Xeroriparian systems are important not only for biodiversity, but the water in them has been a coveted commodity for human survival and productive activities, and, in consequence, they have suffered extreme, widespread modification. As a result, within the past 100 years an estimated 95% of lowland riparian habitat in western North America has been altered, degraded, or destroyed (Krueper, 2000). In arid and semiarid regions where water is naturally scarce, livestock and agricultural demands for it result in riparian systems being affected with particular severity (Patten et al., 2018). Mexico´s arid and semiarid Central Altiplano is no exception, and its riparian systems have been transformed by their water being diverted for human needs with no consideration for the conservation of wildlife (Mellink & Riojas-López, 2005).
Published information on non-urban xeroriparian systems is scarce, and for Mexico, it is even scarcer. The only 3 articles in the scientific literature that we have found on birds in xeroriparian systems in Mexico focus on perennial systems (Arizmendi et al., 2008; Hinojosa-Huerta et al., 2013; Pérez-Amezola et al., 2020). Three highly relevant nationwide biodiversity conservation sources surprisingly do not mention riparian systems: 1) A 1998 listing of Mexico´s natural protected areas (Conabio, 1998); 2) the extensive treatise on the use and conservation of the terrestrial ecosystems of Mexico (Challenger, 1998); and 3) the 3 volume, 1,739 pages, Estado de Mexico´s biodiversity and its conservation threats (Conabio, 2008). This suggests that in addition to being understudied, the importance of riparian systems in Mexico has not been fully appreciated.
One of the least studied landscape components of the southern Mexican Altiplano, including the Llanos de Ojuelos, are xeroriparian systems. This region is strongly anthropized and natural habitats have been greatly affected by agriculture and livestock, including the riparian systems in it. Currently, those riparian systems in the region that have not disappeared because of water channelization and damming, are subject to browsing and trampling by livestock, and by the extraction of wood, sand, gravel, and water. However, the remaining xeroriparian systems in the southern part of this Altiplano, even in their impacted form, continue to provide habitat for wildlife (Riojas-López & Mellink, 2019; Riojas-López et al., 2019).
Birds that use xeroriparian systems in the southern part of the Altiplano are little known, and the limited knowledge about them had so far derived from occasional observations only (for example, in Riojas-López & Mellink, 2019). As pointed out in the literature, intermittent and ephemeral xeroriparian systems are keystone habitats for biodiversity, although their role in Mexico has not been assessed. This keystone role can be expected to be especially important in a country like Mexico where arid and semiarid conditions cover half of its territory (Challenger, 1998). This information void precludes the design of pertinent and timely conservation plans for these habitats and the wildlife that uses them. Hence, in this study we aimed at documenting the birds that use xeroriparian systems in the highly anthropized southern part of the Altiplano, and the habitat characteristics that drive their assemblages during the non-breeding season. We studied 3 xeroriparian systems during the non-breeding bird season, with 2 objectives: 1) document the species richness, abundance and their temporal variation, and 2) determine the relationship between bird species richness and abundance and vegetation characteristics. In the context of an alarming decline of bird populations in North America (Rosenberg et al., 2019), studies like this are needed as a baseline to monitor the trends of bird populations that depend on the persistence of xeroriparian systems. The urgency of this need is increased because of the ongoing climate change in which drier and hotter regimes are predicted.
Materials and methods
The study was carried out in the Llanos de Ojuelos, at the convergence of the states of Jalisco, Zacatecas, Aguascalientes, San Luis Potosí and Guanajuato (Fig. 1). This area is a semi-arid tableland at 1,900-2,600 m altitude with a geomorphology of low mountains and valleys (Nieto-Samaniego et al., 2005). Three climatic seasons occur: dry cold (November-February), dry hot (April-May), and rainy (June to September); March and October are intermediate (Mellink et al., 2016), with an average annual temperature at the Ojuelos de Jalisco, Jalisco, meteorological station (1988-2008) of 15 °C, annual rainfall of 681 mm, and tank evaporation higher than precipitation all months of the year. The area has endorreic drainage, and rainwater flows through ephemeral streams or, in some cases, as sheet flows and collects in seasonal pools or is stored in cattle watering tanks and dams. Historically, springs were common, but the majority have disappeared (Mellink & Riojas-López, 2005).
The natural vegetation of the region is composed of grasslands (42.6% of the region’s surface), xerophilous shrublands (15.66%) and stands of dwarf oaks (Quercus spp.; 4.61%). Grasses of the genera Bouteloua, Aristida, Lycurus, and Mulhenbergia are the most common components of grasslands. Catclaws (Mimosa spp.), silver dalea (Dalea bicolor), leatherstem (Jatropha dioica), huizache (Vachellia spp.), arborescent nopales (Opuntia spp.), Peruvian pepper tree (Schinus molle), and yucca (Yucca spp.) form the shrub and arborescent layers (Harker et al., 2008; MER-L & EM pers. obs.).
Livestock and agriculture are the main productive activities in the Llanos de Ojuelos and have transformed the region´s landscape since the arrival of Spanish conquerors 450 ~ 500 years ago (Mellink & Riojas-López, 2020). Currently, approximately 35.5% of the surface of the municipalities of Ojuelos de Jalisco, Jalisco, and Pinos, Zacatecas is devoted to rain-fed farming of mostly corn, beans and fruit-oriented nopal orchards, while sheep, goats, cows, and horses graze and browse throughout the region (Pers. obs.).
Figure 1. The Llanos de Ojuelos region, southern Mexican Altiplano, indicating the 3 xeroriparian systems where bird communities were studied during the 2019-2020 non-breeding season (in white lettering), reference localities (in green upper/lowercase), and states (in green small caps).
This study was performed in 3 independent and geographically separated xeroriparian systems (localities), through visual surveying of birds during the 2019-2020 non-breeding season (Fig. 1). The localities were selected based on them being safe, accessible, and independent of each other (i.e., that their channels were not connected), and that the owners allowed us to work in them. In each system, we established 3 survey sections along the stream, with different characteristics. These xeroriparian systems, from north to south, were: La Laborcilla (Table 1). Its stream is ephemeral and flows southeast from the low mountain range that stretches between La Montesa and El Nigromante, in the municipality of Pinos. It has a straight riverbed (Sinuosity Index [SI], sensu Rosgen, 1994, < 1.1) and its slope is 8.4%. Boulders cover most of the streambed. The ground in the area adjacent to stream is mostly rocky and covered by shrubland whose major components are junipers (Juniperus deppeana), dwarf oaks, central Mexico yucca, and huisaches. The range is used for the raising of sheep and goats, along with a few cattle. Rancho Santoyo (Table 2). This is a slow-flowing straight stream (SI < 1.1), on sandy and tepetate streambed and low slope (2.4%), with permanent water in parts of it, provided by a permanent spring. The surroundings are overgrazed grassland with huisaches, and shrubland with arboreal nopales, huisaches, and pepper trees. The range is used to raise fighting-bull cattle. La Colorada (Table 3). Draining south from the Mesa del Toro, near Ojuelos, the sandy and tepetate streambed of this system is sinuous (SI = 1.1-1.3), with an overall slope of 2.3%. Its surroundings are of grasslands, some overgrazed and some in good condition, with some huisaches and shrubby nopales, and farmland. The range is used mostly for the raising of fine horses, while the nearby farmland is used to grow beans and chilies.
Table 1
Morphological characteristics and vegetation composition, based on the most common tree and shrub species of different sections of the xeroriparian system of La Laborcilla, in the Llanos de Ojuelos, southern part of the Mexican Altiplano, whose birds were studied during the 2019-2020 non-breeding season. Streambed values are mean ± standard error.
Section | Coordinates | Streambed | Description | |
Lat./Long. | Width (m) | Depth (m) | ||
Upper | 22°5’32”-101°43’33” | 25.8 ± 2.2 | 10.2 ± 2.2 | A deep ravine with dwarf oaks (Quercus spp.) and junipers (Juniperus deppeana), with dispersed maguey (Agave sp.) and sotol plants (Dasylirion spp.). |
Middle | 22°5’26”-101°43’27” | 26.1 ± 0.5 | 10.6 ± 1.0 | A deep ravine, but here the dominant treelike form were junipers and huizaches (Vachellia spp.), with dispersed maguey and sotol plants. |
Lower | 22°5’17”-101°43’4” | 25.0 ± 3.9 | 2.4 ± 0.5 | A shallow ravine whose main arboreal component were junipers, with interspersed yuccas (Yucca spp.), and a shrub layer composed of dispersed leatherstem (Jatropha dioica), jimmyweed (Isocoma spp.), and catclaws (Mimosa spp.). |
We surveyed the birds monthly from September 2019 to March 2020, covering the entire non-breeding season: the migratory seasons of Autumn and Spring, as well as the Winter in-between. Birds were identified and counted for 3 consecutive days at each study system, once per month. Bird inventorying was carried out for 2 hours in the afternoon ending at sunset and 2 hours the following morning starting at sunrise, as these are the periods of highest bird activity. Bird nomenclature and taxonomic arrangement follows Chesser et al. (2023).
Table 2
Morphological characteristics and vegetation composition, based on the most common tree and shrub species of different sections of the xeroriparian system of Rancho Santoyo, in the Llanos de Ojuelos, southern part of the Mexican Altiplano, whose birds were studied during the 2019-2020 non-breeding season. Streambed values are mean ± standard error.
Section | Coordinates | Streambed | Description | |
Lat./Long. | Width (m) | Depth (m) | ||
Upper | 21°55’1”-101°47’32” | 50.9 ± 2.5 | 3.8 ± 0.3 | A relatively wide and moderately deep riverbead, densely vegetated with tall willows (Salix bonplandiana), cottonwoods (Populus fremontii) in addition to pepper trees (Schinus molle) and huizaches, interspersed with patches of ragwort (Senecio spp.). |
Middle | 21°55’10”-101°47’31” | 38.0 ± 2.6 | 1.7 ± 0.5 | A wide and shallow part of the riverbed, covered by peppertrees and ragwort. |
Lower | 21°55’40”-101°47’25” | 12.4 ± 1.5 | 0.7 ± 0.5 | A very thin and shallow canal, flanked by large peppertrees, with huisaches, and some catclaw and low nopales (Opuntia spp.). |
Table 3
Morphological characteristics and vegetation composition, based on the most common tree and shrub species of different sections of the xeroriparian system of La Colorada, in the Llanos de Ojuelos, southern part of the Mexican Altiplano, whose birds were studied during the 2019-2020 non-breeding season. Streambed values are mean ± standard error.
Section | Coordinates | Streambed | Description | |
Lat./Long. | Width (m) | Depth (m) | ||
Upper | 21°47’48”-101°38’18” | 36.1 ± 5.9 | 4.3 ± 0.9 | A deep canyon in which the most notorious trees were dwarf oaks, with dispersed catclaws and sotol, as the most abundant shrubs. |
Middle | 21°47’30”-101°37’34” | 75 ± 5.7 | 6.3 ± 0.8 | Semi-open deep riverbed dominated by pepper trees, with some dispersed huisaches, arborescent nopales, and willows; the most notorious shrubs were ragworts. |
Lower | 21°47’5”-101°36’51” | 61.0 ± 8.3 | 7.0 ± 2.2 | A semi-open riverbed in which the most notorious trees were pepper trees and arborescent nopales, with some huisaches. The most visible shrubs were catclaws, some sedges and some shrubby nopales. |
Three survey stations were established in each of the 3 sections within each xeroriparian system, 40 m apart from each other and each consisting of one survey point in the center of the riverbed and 1 at the edge of the riparian habitat, looking at it. These 2 survey points allowed us to record birds that prefer the outer canopy as well as those that prefer the understory. One “day” of observation consisted of an afternoon and the following morning. We randomized the order in which the survey points were surveyed in such a way that each section was surveyed on 1 day in the first place, on 1 day in the second place, and on 1 day in the third place. For example, on day 1 survey order could be B1 [section B, station 1], A2, C3; on day 2 C1, B2, A1; and on day 3: A3, C2, B3. In some cases, riverbed survey points became darker sooner in the afternoon and lighter in the morning. Therefore, the riverbed points were surveyed before their corresponding outside station in the afternoon, and after it in the morning. Within a survey month, the same randomization was applied to the 3 xeroriparian systems, but a new randomization was performed every month. Survey order of xeroriparian systems followed logistic considerations and varied between months.
Birds at each survey point were identified and counted with 8×40 binoculars in a circle with a 20-m radius for 10 min (Brand et al., 2008; Merrit & Bateman, 2012). We did not include the birds that were observed outside or flying over the riparian system. As the best proxy of each species’ abundance in any given station we selected the highest count among the 4 counts carried out on a visit: riverbed and outside survey points, afternoon, and morning counts (Merrit & Bateman, 2012). For each section, the monthly estimate of abundance for each species was the sum of the 3 stations’ maximum values. Trophic guild and residency status of each bird species were obtained from the “The Birds of the World” series of monographs (https://birdsoftheworld.org/bow/home).
Additional to bird surveys, we measured vegetation attributes that according to the literature are important for birds (Brand et al., 2008; Powell & Steidl, 2015; Rotenberry, 1985; Wiens & Rotenberry, 1981). On 1 visit, we identified the dominant plants and measured the minimum and maximum height of shrubs and of trees. On each survey period, we determined herb cover of the ground, herb vertical density, and canopy cover. Herb cover of the ground in the region grows explosively as a result of Summer rains and dries after maturation. We used a simple scoring of 3 levels of ground cover by herbs: completely bare or nearly so, medium cover, and completely covered, or nearly so based on visual appreciation.
Herb vertical density was calculated with a 30-cm wide board divided into bands every 25 cm in height until 100 cm (0-25, 25-50, 50-75 and 75-100 cm). This board was placed 10 m away from each internal survey point in 4 directions, 2 parallel and 2 perpendicular to the streambed, and the percentage of visual obstruction in each band as seen from the center point was recorded (Hays et al., 1981).
Canopy cover was determined by foliage cover in 4 photographs of the canopy with an inclination of 30° from the vertical, at all streambed points. Two photographs were taken along the stream axis and 2 perpendicular to it, 1 to each side. The percentage of obstruction of the vegetation in each photograph was calculated counting number of pixels with and without vegetation using Photoshop ver. 2017.
Through an information-theoretic approach (Burnham & Anderson, 2002), we tested the effect of xeroriparian system and season (fixed effects) on richness, overall abundance, and abundance of the bird species that summed more than 10 individuals. Poisson distribution was used, and survey section was included as a random effect. We selected the best model with Akaike Information Criterion for small samples (AICc), applying the principle of parsimony when differences in AICc values were < 2.5 (Burnham & Anderson, 2002). Whenever we refer to a “best model” it implies that it was either the best or the most parsimonious model. We also used the same approach to explore the influence of vegetation attributes on the same bird variables. Before running the models, we obtained correlations between habitat variables and averaged those that were correlated > 0.85 and reviewed the new correlation values. This procedure was used to prevent any important variable for the birds from going unnoticed by not being part of the best model as a result of the variance that it would explain being partially accounted for by another, highly correlated variable which was included in such model. The final list of explanatory habitat variables included density of herbs, visual obstruction at 0-0.25 cm, 25-75 cm and 75-100 cm, minimum height of shrubs, minimum height of trees, mean and maximum height of shrubs and trees averaged, and canopy cover. Modeling was performed using pgirmess and lme4 libraries in R 3.3.1, through RStudio Ver. 1.2.5019.
Averaged vegetation characteristics were compared between systems through analyses of variance, followed by post-hoc Tuckey tests if statistical differences were detected (p ≤ 0.05). Similarity between systems was calculated through Jaccard´s index. We arranged study sections through Principal Components Analysis (PCA) based on their birds, both on binary data (species presence/absence) and on their abundance. These analyses were done using PAST 4.03 (https://folk.universitetetioslo.no/ohammer/past).
Results
During our surveys we identified 48 species of birds, in addition to some individuals that were identified only at genus or family level, but which likely belonged to 1 of the identified species (Supplementary material 1, 2). The identified species included 30 resident and 18 migratory bird species. Twelve additional species were recorded using xeroriparian systems, but outside our surveys. The total count during surveys was 932 individuals. Spizella passerina was the most abundant species with 149 individuals (16% of total abundance). Five other species contributed between 5 and 10% of the total abundance: Corthylio calendula, Setophaga coronata, Zenaida asiatica, Psaltriparus minimus, and Aphelocoma woodhouseii (Supplementary material 2).
Figure 2. Overall abundance (number of individuals) of bird guilds in the Llanos de Ojuelos region, southern Mexican Altiplano, during the 2019-2020 non-breeding season in 3 xeroriparian systems.
Figure 3. Monthly abundance of all birds and birds of resident species that included survey month as part of best models exploring the influence of locality (xeroriparian system) and month, during the 2019-2020 non-breeding season in 3 xeroriparian systems in the Llanos de Ojuelos, southern Mexican Altiplano.
Figure 4. Monthly abundance of birds of migrant species that included survey month as part of best models exploring the influence of locality (xeroriparian system) and month, during the 2019-2020 non-breeding season in 3 xeroriparian systems in the Llanos de Ojuelos, southern Mexican Altiplano.
Figure 5. Total abundance of species that had location as part of the best model exploring the influence of locality and survey month, during the 2019-2020 non-breeding season in each of 3 xeroriparian systems in the Llanos de Ojuelos, southern Mexican Altiplano.
La Laborcilla had a total bird count of 213 individuals (162 of 13 resident/breeding species, 48 of 8 migrant species, and 3 of 2 unidentified species), Rancho Santoyo had 353 (197/22, 149/11, and 7/3), and La Colorada, 366 (228/27, 128/16, and 10/3). Most individuals counted were of resident species or of 1 Summer (breeding) resident species (Myiarchus cinerascens),while migrants were a smaller component of the community (Fig. 2). For species richness, the best model did not include locality nor month. The best model explaining overall bird abundance included month (Supplementary material 3), but not locality. Abundance increased from Autumn (September) to Spring (March) with some differences between guilds. Insectivore migrants increased in number into the Winter and then decreased, while granivore migrants began to arrive in December and increased until March (Fig. 2). In all these cases the month of survey was part of the best model, as it was of 11 bird species (Figs. 3, 4; Supplementary material 3). Locality was part of the best model explaining abundance in 4 cases out of 27, all of them individual species (Fig. 5; Supplementary material 3). Aphelocoma woodhouseii and Leiothlypis celata were the only 2 species whose best model included both locality and moth, while the best models for the other species did not include either variable (Supplementary material 3).
Figure 6. Total abundance of birds in different trophic guilds in 3 xeroriparian systems and 3 sections within each, during the 2019-2020 non-breeding season in the Llanos de Ojuelos, southern Mexican Altiplano. Numbers on the circles indicate richness and total abundance, while numbers on the side of the dendrograph indicate similarity between xeroriparian systems, Bird assemblages in Rancho Santoyo and La Colorada were more similar between them than to La Laborcilla (Fig. 6). In PCA graphs, the 3 sections at La Laborcilla grouped closer to each other than those of the other systems. The sections within each system grouped discreetly when binary data was used (Fig. 7 top), but groups overlapped when based on abundance (Fig. 7 bottom). The 3 systems studied differed in the resident/breeding vs. migrant composition of the communities, and sections within systems were also different (Fig. 6). Whereas Rancho Santoyo had a higher proportion of insectivore migrants than the other locations, La Colorada had a higher count of granivore migrants, and La Laborcilla had proportionally more individuals of resident species (Myiarchus cinerascens did not occur in La Laborcilla). In neither case did such patterns occur in the 3 sections of the corresponding system, but only in 2 of Rancho Santoyo’s sections and in 1 and partially in another at La Colorada.
Figure 7. Principal Component arrangement of 3 study sections in each of 3 xeroriparian systems based on their bird composition during the 2019-2020 non-breeding season in the Llanos de Ojuelos, southern Mexican Altiplano. The upper figure is based on binary data (presence/absence of species), and the lower figure, on bird abundance. Ellipses were drawn around the 3 sections of each system by hand. The legends “Up”, “Md”, and “Lw” indicate the upper, middle and lower sections of each system.
The attributes of the plant communities were different between xeroriparian systems (Table 4). La Laborcilla had significantly less ground covered by herbs, whereas Rancho Santoyo and La Colorada were not different in this aspect. La Colorada had significantly denser vegetation from 0 to 75 cm above the ground than the 2 other systems, but visual obstruction at 75-100 cm was not different between the 3 systems. The tallest shrubs and highest trees were significantly taller in Rancho Santoyo than in La Laborcilla, whereas La Colorada was not different from either, and average height both shrubs and trees was significantly greater at Rancho Santoyo than at La Laborcilla, with La Colorada being in between and statistically different from either. Canopy cover was not different between systems. Study sections were all peculiar within the systems, but only in some cases were sections significantly different (Supplementary material 4). Despite slight differences, visual obstruction at the 4 heights assessed was highest in September and October, and then decreased towards their lowest values in March (Fig. 8).
At least 1 habitat attribute was part of the best bird model in all but 3 cases (Tables 5-7), the 3 of them resident species: Zenaida asiatica, Thryomanes bewickii, and Phainopepla nitens. In all cases in which canopy cover was part of the best model, it had a negative effect on bird richness or abundance, while herb density and visual obstruction at 25-75 cm had a negative effect in most of the best models that included them (Tables 5-7). In contrast, visual obstruction at 75-100 cm and mean/maximum height of shrubs/trees had a positive effect.
Figure 8. Mean visual obstruction in 4 25-cm vertical layers, between 0 and 100 cm above the ground across 3 sections in each of 3 xeroriparian systems studied during the 2019-2020 non-breeding season in the Llanos de Ojuelos, southern Mexican Altiplano. To reflect the spatial arrangement of the information, the panels are arranged bottom to top. The height stratum to which each graph corresponds is indicated on the graph. Points on each graph with the same letter are not statistically different (p < 0.05).
Discussion
The lack of studies about the role of ephemeral and intermittent xeroriparian systems as key habitats for biodiversity conservation and potential provision of ecological services severely impairs the capability of designing and implementing timely and informed conservation actions. In this study we generated a basic understanding about the composition of bird assemblages in 3 Mexican xeroriparian systems and the habitat attributes that influence them. Our study is particularly pertinent as the results of only 3 other research projects on birds in Mexican xeroriparian systems have been published (Arizmendi et al., 2008; Hinojosa-Huerta et al., 2013; Pérez-Amezola et al., 2020), none of them including ephemeral systems.
The 48 terrestrial species that we recorded represent 20% of all potential terrestrial native birds of the area, excluding swifts (family Apodidae) (based on Howell and Webb [1995]). This is relevant considering that xeroriparian systems, in general, occupy around 5% of the total land surface in western North America (Krueper, 2000), and that those we studied covered approximately 0.1% of the area in which they are located (by delineating them in Google Earth and measuring their area as well as that of the displayed image). According to Partners of Flight (2023) Campylorhynchus brunneicapillus, Spizella atrogularis, Selasphorus rufus/sasin, and Cardellina pusilla have conservation problems, while Accipiter striatus and A. cooperii are protected by Mexican law (Semarnat, 2010). The importance of xeroriparian systems in the region studied is enhanced by the fact that some otherwise woodland bird species largely depend on them, and we have not documented them in any other xeric habitats of the region (Mellink et al., 2016, 2017; Riojas-López & Mellink, 2019; Riojas-López et al., 2019).
Table 4
Vegetation attributes of 3 xeroriparian systems in the Llanos de Ojuelos, southern part of the Mexican Plateau, 2019-2020. Values are estimate ± standard error, except on maximum and minimum heights. Values of any variable with different literal were significantly different (p ≤ 0.05) according to an ANOVA + Tukey post hoc tests. Superscript “ns” indicates that the means were not significantly different.
Site/section | Herb | Visual obstruction (%) | Hight of shrubs (m) | Height of trees (m) | Canopy | |||||||
density (1-4) | 0-25 cm | 25-50 cm | 50-75 cm | 75-100 cm | Max | Min | Mean | Max | Min | Mean | cover (%) | |
La Laborcilla | 1.53 ± 0.38b | 12.00 ± 4.27b | 4.22 ± 2.58b | 0.89 ± 1.06b | 0.40 ± 0.55ns | 2.63 ± 0.32b | 0.23 ± 0.10ns | 0.63 ± 0.01c | 8.37 ± 0.27b | 0.73 ± 0.26ns | 3.58 ± 0.26c | 40.40 ± 8.82ns |
Rancho Santoyo | 2.52 ± 0.88a | 13.60 ± 16.7b | 3.03 ± 4.75b | 0.85 ± 2.38b | 0.83 ± 2.27ns | 3.55 ± 0.49a | 0.31 ± 0.08ns | 1.41 ± 0.31a | 14.65 ± 3.04a | 1.20 ± 0.44ns | 4.79 ± 0.89a | 39.50 ± 16.83ns |
La Colorada | 2.70 ± 0.95a | 35.98 ± 17.06a | 16.04 ± 10.83a | 6.40 ± 6.32a | 1.59 ± 2.48ns | 3.14 ±0.76ab | 0.22 ± 0.06ns | 1.21 ± 0.44b | 10.36 ± 3.55ab | 0.87 ± 0.19ns | 4.54 ± 0.54b | 45.48 ±29.86ns |
These species can be considered locally xeroriparian-dependent. They are Accipiter striatus, A. cooperii, Empidonax difficilis occidentalis, Pitangus sulphuratus, Empidonax wrightii, Corthylio calendula, Turdus migratorius, Setophaga coronata, Setophaga townsendi, Piranga ludoviciana, Bubo virginianus, Sphyrapicus varius, and Cardinalis cardinalis. This was also the case of Pipilo maculatus and P. chlorurus, otherwise species of thick shrublands. The system with more xeroriparian-dependent species was La Colorada (9 species) followed by Santoyo (7) and La Laborcilla (6). However, Santoyo had more riparian-dependent individuals (139) than La Colorada (78) and La Laborcilla (45), Setophaga coronata and Corthylio calendula being the most abundant species (Supplementary material 2).
Table 5
Sign of the effects of habitat features on the bird community variables in 3 study sections in each of 3 xeroriparian systems studied during the 2019-2020 non-breeding season in the Llanos de Ojuelos, southern Mexican Altiplano. The only data indicated are that of variables included in the best or more parsimonious model under an information-theoretic approach. Blank cells are of variables not included in such models. Min. indicates minimum, and Max., maximum. Variables with correlation coefficients > 0.85 were merged before the analysis. The intercept is not shown. Actual data is presented in Supplementary material 5.
Bird community variable | Herb | Visual obstruction (%) | Shrub Min. | Tree Min. | Shrub-tree | Canopy | ||
Density (1-4) | 0-0.25 cm | 25-75 cm | 75-100 cm | Height (m) | Height (m) | Min./Max. (m) | Cover (%) | |
Richness | + | – | ||||||
Abundance | ||||||||
Overall | – | – | + | – | ||||
All resident species | – | – | – | |||||
All migrants | – | + | + | |||||
Migrant insectivorous birds | – | + | – | |||||
Migrant granivorous birds | – | + | – |
Table 6
Sign of the effects of habitat features on the abundance of resident species of birds in 3 study sections in each of 3 xeroriparian systems studied during the 2019-2020 non-breeding season in the Llanos de Ojuelos, southern Mexican Altiplano. The only data indicated are that of variables included in the best or more parsimonious model, under an information-theoretic approach. Blank cells are of variables not included in such models. Min. indicates minimum, and Max., maximum. Variables with correlation coefficients > 0.85 were merged before the analysis. The intercept is not shown. Actual data is presented in Supplementary material 5.
Bird response variable | Herb | Visual obstruction (%) | Shrub Min. | Tree Min. | Shrub-tree | Canopy | ||
Density (1-4) | 0-0.25 cm | 25-75 cm | 75-100 cm | Height (m) | Height (m) | Min./Max. (m) | Cover (%) | |
Zenaida asiatica | ||||||||
Melanerpes aurifrons | – | |||||||
Sayornis nigricans | + | |||||||
Aphelocoma woodhouseii | – | – | ||||||
Psaltriparus minimus | – | + | – | – | ||||
Phainopepla nitens | ||||||||
Thryomanes bewickii | ||||||||
Mimus polyglottos | + | |||||||
Spinus psaltria | + | + | + | – | ||||
Spizella passerina | – | – | – | |||||
Melozone fusca | + | |||||||
Pipilo maculatus | – | + | – |
Table 7
Sign of the effects of habitat features on the abundance of migratory birds in 3 study sections in each of 3 xeroriparian systems studied during the 2019-2020 non-breeding season in the Llanos de Ojuelos, southern Mexican Altiplano. The only data indicated are that of variables included in the best or more parsimonious model, under an information-theoretic approach. Blank cells are of variables not included in such models. Min. indicates minimum, and Max., maximum. Variables with correlation coefficients > 0.85 were merged before the analysis. The intercept is not shown. Actual data is presented in Supplementary material 5.
Bird response variable | Herb | Visual obstruction (%) | Shrub Min. | Tree Min. | Shrub-tree | Canopy | ||
Density (1-4) | 0-0.25 cm | 25-75 cm | 75-100 cm | Height (m) | Height (m) | Min./Max. (m) | Cover (%) | |
Empidonax wrightii | + | |||||||
Corthylio calendula | – | + | ||||||
Troglodytes aedon | + | |||||||
Turdus migratorius | – | – | + | – | ||||
Melospiza lincolnii | – | – | + | – | ||||
Spizella pallida | – | + | – | + | – | |||
Leiothlypis celata | + | |||||||
Setophaga coronata | – | |||||||
Cardellina pusilla | + |
Resident species increased their abundance from December through March (Figs. 2, 3), a pattern that could have been driven by 3 processes, not necessarily mutually exclusive: 1) resident species that might breed in xeroriparian habitats disperse to feed in other habitats in the region after nesting, might have begun to congregate for the upcoming breeding season, which would make the different habitats complementary (Dunning et al., 1992); 2) species like the P. maculatus might become more detectable as breeding-associated territoriality and courting behaviors develop; 3) migrating individuals of northern populations of species that locally remain resident might pass through the region in the Spring (perhaps S. passerina; Fig. 4).
The increase in abundance of the resident species as the season progressed was combined with the addition of migratory species in their northbound Winter-Spring migration. The pattern observed in our study is similar to that in the lower Colorado River in southwestern Arizona, where more birds migrate during the Spring than during the Autumn and suggests that some species migrate via different routes and/or use different habitats during the latter (Carlisle et al., 2009). According to our data, the region is part of the Spring migration route, but not, or less so, of the Autumn route.
Birds assemble differently in response to habitat characteristics (Wiens & Rotenberry, 1981), although bird-habitat relationships are complex (Strong & Bock, 1990). In our study, habitat characteristics were part of the best model in 78% of the cases (Tables 5-7). Three habitat variables more consistently affected bird assemblages and species. The first was that the averaged minimum and maximum height of shrubs and trees influenced birds positively, which conforms with general known principles in bird ecology (Brand et al., 2008; MacArthur & MacArthur, 1961; Merrit & Bateman, 2012; Rockwell & Stephen, 2018), and with findings in other xeroriparian systems (Brand et al., 2008). The second was that closed canopies had a negative effect. Closed canopies have been found to influence birds negatively by reducing light at ground level resulting in a less developed herb and shrub community (Beedy, 1981). The third case was that herb vertical density at 25-75 cm had also a negative effect. But this is a spurious outcome that resulted from the dense herb layer at the beginning of the study (Fig. 8), when bird abundance variables were lower, while the circannual process of herbs drying and decaying in the late Autumn and early Winter causes low herb cover coinciding with the increase in bird abundance (Figs. 2, 3).
The bird assemblages of the xeroriparian systems studied not only were different from each other, but also varied internally, between sections (Figs. 6, 7). The arrangement of the 9 sections according to their species presence/absence in axis 1, explaining 33% of the variance, follows a clear gradient (Fig. 7 top). On the left side are sections with steep, narrow, and rocky ravines with large boulders, dominated by oaks and junipers that form a close canopy with little understory herbaceous vegetation. On the right, wide and open pebbled washes, dominated by a more heterogeneous tree community composed of peppertrees, arborescent nopales, and huisaches, with a well-developed herbaceous layer. The bird communities responded to this gradient with increasing richness and abundance from left to right.
In the same graph, axis 2 follows a gradient from a well-developed shrub community and abundant litter and wetter soil to no or scarce shrubs with a dense herbaceous layer and drier ground. This gradient apparently drives the food resources available to birds, as suggested by the preponderant bird guilds in the different sections (Fig. 6): insects in its lower end to seeds in the upper end. Overall, La Colorada with the most heterogeneous sections, vegetation-wise, had also the richest bird assemblage, while La Laborcilla, with the most homogenous sections, had the poorest one, with Rancho Santoyo intermediate in both attributes (Table 4; Fig. 6; Supplementary material 5). Using bird abundance instead of binary information rearranged the PCA graph layout (Fig. 7 lower) because of the effect of species that were widespread and abundant, but without losing a resemblance to the species presence/absence PCA.
Vertical structure is important for birds, but floristic composition can also influence their diversity (Fleishman et al., 2003; Rotenberry, 1985). For example, oaks are of little attraction to most birds (Powell & Steidl, 2015), and their dominance at La Laborcilla and the upper section of La Colorada coincided with the lower richness and abundance of their bird assemblages (Fig. 6; Supplementary material 2). Rancho Santoyo supported more xeroriparian-dependent individuals, mostly of Setophaga coronata and Corthylio calendula, both small insectivorous birds. The middle and upper sections where these 2 species thrived had large cottonwoods (Populus fremontii) and willows (Salix bonplandiana). In contrast, in the middle section of La Colorada, which also had tall trees but were peppertrees and arborescent nopales, these 2 bird species were much less abundant.
Each of the systems studied had a distinct signature given by one species or dominant bird guild(s), although such signature was largely due to the assemblages in the individual sections (Fig. 6). The signature species at Laborcilla was Aphelocoma woodhouseii, whose primary habitat includes oak and juniper forests, where it typically feeds on juniper berries in the Autumn and Winter (Fig. 5; Cornell Lab of Ornithology, 2019). Although the upper Colorada section grouped with La Laborcilla, with which it shared oaks, lacked junipers and was not used by Aphelocoma woodhouseii. Their close PCA grouping rather resulted from their shared poor bird assemblages. Rancho Santoyo can be identified with migrant insectivorous birds and La Colorada with migrant granivorous birds, in concordance with the explanation on the PCA arrangement provided above.
The lower proportion and abundance of migratory insectivorous birds and absence of granivorous birds at La Laborcilla might have been caused by its less developed understory vegetation and enclosed canyon conditions. In contrast, Rancho Santoyo´s upper and middle section provided the best habitat for insectivore migrants. This system was exuberant in herb and shrub foliage, likely as a result of the longer presence of ground humidity, but had also plenty of sunny spots; and the system bordered open rangeland, providing lengthy, sharp borders, especially in its upper and middle sections. Granivore migrants were a small component of the communities we documented, and they were present almost exclusively in La Colorada´s middle section. This site had an open canopy and combined taller shrubs with denser and taller herbs than in the other sections of the same and the 2 other systems, providing higher habitat heterogeneity and, likely, more seeds. A more developed herb-shrub stratum at 0-75 cm at La Colorada than at other sites provided good escape cover adjacent to open patches that provided seeds (Table 4; Supplementary material 4).
As our data exhibit, habitats associated with non-perennial xeroriparian streams are far from uniform, not only between systems but also within them. At a landscape level they are clearly keystone structures, at least for the birds but surely for other groups as well, both collectively and individually. The 3 systems studied by us cover about 55 ha, roughly 0.1% of the area in which they occur. However, they supported 20% of the potential species of terrestrial birds of the region, including 15 that we have recorded only in such riparian habitat. Our data exhibit that in addition to their importance for resident species, some ephemeral and intermittent xeroriparian habitat in the southern part of the Mexican Altiplano are important for northbound Spring migrating birds.
Our study is a first approximation to the ecological role of xeroriparian systems in the region. However, many issues, like their importance as nesting habitat, provision of food, climatic protection, interaction with adjacent and farther away habitats, among others, remain to be studied. Nevertheless, if these xeroriparian habitats disappear, regional biodiversity would be impacted. Not only 1 or a few, but many or all xeroriparian systems, and their different sections in the region studied by us should be targeted for conservation management. Xeroriparian systems have long been considered important elements of the landscape and their conservation needs recognized. However, such consideration, and the actions derived from it, usually focus on perennial streams with their lush arboreal communities. This focus is biased and excludes an important part of xeroriparian habitat: that created by ephemeral and intermittent streams. Despite their low consideration in research and in conservation agendas, as our study and a few others have demonstrated, ephemeral and intermittent xeroriparian habitat can play a crucial role in arid and semiarid lands (Johnson & Haight, 1985; Levick et al., 2008; Sánchez-Montoya et al., 2017; Szaro & Jakle, 1985). Despite covering less than 0.1% of the region´s area, in our study they supported 20% of all terrestrial species that we documented in the region (Mellink et al., 2016, 2017; Riojas-López & Mellink, 2019; Riojas-López et al., 2019), while those that we documented only in xeroriparian systems account for 9% of all species documented. A scenario of ecological importance of non-perennial xeroriparian systems and research and management neglect are likely to occur in many other arid and semiarid regions of the world. Hence, it might be time to join forces and impulse a global agenda for their conservation, which is now especially pertinent in view of the ongoing climate change in which drier and hotter regimes are predicted.
Acknowledgements
Jaime Luévano Esparza, David H. Almanzor, Santiago Cortés, and Marco A. Carrasco assisted during field work. Access and research permission were granted kindly by owners Family Santoyo (Rancho Santoyo) and Enrique Campos (La Laborcilla), and ranch manager Melquíades Contreras (La Colorada). Ezequiel Martínez and Margarita Chávez provided logistic support. Two anonymous reviewers provided extensive and valuable comments. Our greatest appreciation to all of them. Financial support was provided by the Centro de Investigación Científica y de Educación Superior de Ensenada (CICESE), the Universidad de Guadalajara, and the first two authors´ personal funds. The Consejo Nacional de Humanidades, Ciencias y Tecnologías supported MM-O through a M.Sc. scholarship.
References
Arizmendi, M. D., Dávila, P., Estrada, A., Figueroa, E., Márquez-Valdelamar, L., Lira, R. et al. (2008). Riparian mesquite bushes are important for bird conservation in tropical arid Mexico. Journal of Arid Environments, 72,1146–1163. https://doi.org/10.1016/j.jaridenv.2007.12.017
Beedy, E. C. (1981). Bird communities and forest structure in the Sierra Nevada of California. Condor, 83,97–105. https://doi.org/10.2307/1367415
Brand, L. A., White, G. C., & Noon, B. R. (2008). Factors influencing species richness and community composition of breeding birds in a desert riparian corridor. Condor, 110,199–210. https://doi.org/10.1525/cond.2008.8421
Burnham, K. P., & Anderson, D. R. (2002). Model selection and multimodel inference: A practical information-theoretic approach. 2nd Ed. New York: Springer-Verlag.
Carlisle, J. D., Skagen, S. K., Kus, B. E., Van Riper III, C., Paxtons, K. L., & Kelly, J. F. (2009). Landbird migration in the American West: recent progress and future research directions. Condor, 111,211–225. https://doi.org/10.1525/cond.2009.080096
Challenger, A. (1998). Utilización y conservación de los ecosistemas terrestres de México. México D.F.: Comisión Nacional para el Conocimiento y Uso de la Biodiversidad, Instituto de Biología (UNAM), and Agrupación Sierra Madre.
Chesser, R. T., Billerman, S. M., Burns, K. J., Cicero, C., J. L. Dunn, J. L., Hernández-Baños, B. E. et al. (2023). Check-list of North American Birds (online). American Ornithological Society. Retrieved on March 14th, 2024 from: https://checklist.americanornithology.org/taxa/
Conabio (Comisión Nacional para el Conocimiento y Uso de la Biodiversidad). (1998). La diversidad biológica de México: estudio de país. México D.F.: Comisión Nacional para el Conocimiento y Uso de la Biodiversidad.
Conabio (Comisión Nacional para el Conocimiento y Uso de la Biodiversidad). (2008). Capital natural de México, vol. I: Conocimiento actual de la biodiversidad. México D.F.: Comisión Nacional para el Conocimiento y Uso de la Biodiversidad.
Cornell Lab of Ornithology (2019). Woodhouse´s Scrub-Jay. Retrieved on June 7th, 2023 from: https://www.allaboutbirds.org/guide/Woodhouses_Scrub-Jay/lifehistory#
Datry, T., Bonada, N., & Boulton, A. J. (2017). Conclusions: recent advances and future prospects in the ecology and management of intermittent rivers and ephemeral streams. In T. Datry, N. Bonada, & A. J. Boulton (Eds.), Intermittent rivers and ephemeral streams: ecology and management (pp. 563–584). San Diego, CA: Academic Press.
Dunning, J. B., Danielson, B. J., & Pulliam, H. R. (1992). Ecological processes that affect populations in complex landscapes. Oikos, 65,169–175. https://doi.org/10.2307/3544901
Fleishman, E., McDonal, N., Nally, R. M., Murphy, D. D., Walters, J., & Floyd, T. (2003). Effects of floristics, physiognomy and non-native vegetation on riparian bird communities in a Mojave Desert watershed. Journal of Animal Ecology, 72, 484–490. https://doi.org/10.1046/j.1365-2656.2003.00718.x
Harker, M., García R., L. A., & Riojas-López, M. E. (2008). Composición florística de cuatro hábitats en el rancho Las Papas de Arriba, municipio de Ojuelos de Jalisco, Jalisco, México. Acta Botanica Mexicana, 85,1–29.
Hays, R. L., Summers, C., & Seitz, W. L. (1981). Estimating wildlife habitat variables. Washington DC: U.S. Fish and Wildlife Service.
Hinojosa-Huerta, O., Soto-Montoya, E., Gómez-Sapiens, M., Calvo-Fonseca, A., Guzmán-Olachea, R., Butrón-Méndez, J. et al. (2013). The birds of the Ciénega de Santa Clara, a wetland of international importance within the Colorado River delta. Ecological Engineering, 59,61–73. https://doi.org/10.1016/j.ecoleng.2013.03.005
Johnson, R. R., & Haight, L. T. (1985). Avian use of xeroriparian ecosystems in the North American warm deserts. In R. R. Johnson, C. D. Ziebell, D. R. Patton, P. F. Ffolliott, P. F., & R. H. Hamre (Tech. Coords.), Riparian ecosystems and their management: reconciling conflicting uses (pp. 156–160). General Technical Report RM-GTR-120. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station.
Kirkpatrick, C., Conway, C., & LaRoche, D. (2009). Surface water depletion and riparian birds. Tucson, Arizona: Arizona Cooperative Fish and Wildlife Research Unit.
Knopf, F. L. (1985). Significance of riparian vegetation to breeding birds across an altitudinal cline. In R. R. Johnson, C. D. Ziebell, D. R. Patton, P. F. Ffolliott, & R. H. Hamre (Tech. Coords.), Riparian ecosystems and their management: reconciling conflicting uses (pp. 105–111). General Technical Report RM-GTR-120. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station.
Knopf, F. L., Johnson, R. R., Rich, T., Samson, F. B., & Szaro, R. C., 1988. Conservation of riparian ecosystems in the United States. Wilson Bulletin, 100,272–284. http://www.jstor.org/stable/4162566
Krueper, D. J. (1993). Effects of land use practices on western riparian ecosystems. In D. M. Finch, & P. W. Stangel (Eds.), Status and management of neotropical migratory birds (pp. 321–330). General Technical Report RM-229. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station.
Krueper, D. J. (1996). Effects of livestock management on southwestern riparian ecosystems. In D. W. Shaw, & D. M. Finch (Tec. Coords.), Desired future conditions for Southwestern riparian ecosystems: bringing interests and concerns together (pp. 281–301). General Technical Report RM-GTR-272. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station.
Krueper, D. J. (2000). Conservation priorities in naturally fragmented and human-altered riparian habitats of the arid West. In R. Bonney, D. N. Pachley, R. J. Cooper, & L. Nioes (Eds.), Strategies for bird conservation: the partners in flight planning process (pp. 88–90). USDA Forest Service Proceedings RMRS-P-16. Ogden, Utah, U.S.A.: Rocky Mountain Research Station.
Levick, L., Fonseca, J, Goodrich, D., Hernández, M., Semmens, D., Stromberg, J. et al. (2008). The ecological and hydrological significance of ephemeral and intermittent streams in the arid and semi-arid American southwest. Washington D.C.: Office of Research and Development, U.S. Environmental Protection Agency.
MacArthur R. H., & MacArthur, J. W. (1961). On bird species diversity. Ecology, 423,594–598. https://doi.org/10.2307/1932254
McDonough, O. T., Hosen, J. D., & Palmer, M. A. (2011). Temporary streams: the hydrology, geography, and ecology of non-perennially flowing waters. In H. S. Elliot, & L. E. Martin (Eds.), River ecosystems: dynamics, management and conservation (pp. 259–289). Hauppauge, N.Y.: Nova Science.
Mellink, E., & Riojas-López, M. E. (2020). Livestock and grassland interrelationship along five centuries of ranching the semiarid grasslands on the southern highlands of the Mexican Altiplano. Elementa Science of the Anthropocene, 8, 20. https://doi.org/10.1525/elementa.416
Mellink, E., Riojas-López, M. E., & Giraudoux, P. (2016). A neglected opportunity for bird conservation: the value of a perennial, semiarid agroecosystem in the Llanos de Ojuelos, central Mexico. Journal of Arid Environments, 124,1–9. https://doi.org/10.1016/j.jaridenv.2015.07.005
Mellink E., Riojas-López, M. E., & Cárdenas-García, M. (2017). Biodiversity conservation in an anthropized landscape: Trees, not patch size drive bird community composition in a low-input agroecosystem. Plos One, 12, e0179438. https://doi.org/10.1371/journal.pone.0179438
Merritt, D. M., & Bateman, H. L. (2012). Linking stream flow and groundwater to avian habitat in a desert riparian system. Ecological Applications, 22, 1973–1988. https://doi.org/10.1890/12-0303.1
Neate-Clegg, M. H. C., Horns, J. J., Buchert, M., Pope, T. L., Norvell, R., Parrish, J. R. et al. (2021). The effects of climate change and fluctuations on the riparian bird communities of the arid Intermountain West. Animal Conservation, 25,325–341. https://doi.org/10.1111/acv.12755
Nieto-Samaniego, Á. F., Alaniz-Ávarez, S. A., & Camprubí, A. (2005). La Mesa Central de México: estratigrafía, estructura y evolución tectónica cenozoica. Boletín de la Sociedad Geológica Mexicana, 57, 285–318. https://doi.org/10.18268/bsgm2005v57n3a3
Partners in Flight Databases. (2023). Retrieved on July 13th,
2023 from: https://pif.birdconservancy.org/avian-conserva
tion-assessment-database-scores/
Patten, D. T., Carothers, S. W., Johnson, R. R., & Hamre, R. H. (2018). Development of the science of riparian ecology in the semi-arid western United States. In R. R. Johnson, S. W. Carothers, D. M. Finch, K. J. Kingsley, & J. T. Stanley (Tech. Eds.), Riparian research and management: past, present, future, Volume 1 (pp. 1–16). General Technical Report RM-GTR-373. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station.
Pérez-Amezola, M. C., Gatica-Colima, A. B. M. Cuevas-Ortalejo, D. M., Martínez-Calderas, J. M., & Vital-García, C. (2020). Riparian biota of the Protected Area of Flora and Fauna Santa Elena canyon, Mexico. Revista Bio Ciencias, 7, e798. https://doi.org/10.15741/revbio.07.e798
Powell, B. F., & Steidl, R. J. (2015). Influence of vegetation on montane riparian bird communities in the sky islands of Arizona, USA. Southwestern Naturalist 60,65–71. https://doi.org/10.1894/MCG-09.1
Riojas-López, M. E., & Mellink, E. (2019). Registros relevantes de aves en el sur del Altiplano Mexicano. Huitzil, 20, e-513. https://doi.org/10.28947/hrmo.2019.20.2.457
Riojas-López, M. E., Mellink, E., & Almanzor-Rojas, D.H. (2019). Estado del conocimiento de los carnívoros nativos (Mammalia) en un paisaje antropizado del Altiplano Mexi-
cano: el caso de Los Llanos de Ojuelos. Revista Mexicana de Biodiversidad, 90,e902669. https://doi.org/10.22201/ib.
20078706e.2019.90.2669
Riparian Habitat Joint Venture (2004). The riparian bird conser-
vation plan: a strategy for reversing the decline of riparian associated birds in California. Ver. 2.0. Retrieved on September
28th, 2021 from: https://web.archive.org/web/2018072200
1457id_/http://www.prbo.org/calpif/pdfs/riparian_v-2.pdf
Rockwell, S. M., & Stephens, J. L. (2018). Habitat selection of riparian birds at restoration sites along the Trinity River, California. Restoration Ecology, 26,767–777. https://doi.org/10.1111/rec.12624
Rosenberg, K. V, Dokter, A. M., Blancher, P. J., Sauer, J. R., Smith, A. C., Smith, P. A. et al. (2019). Decline of the North American avifauna. Science, 366,120–124. https://doi.org/10.1126/science.aaw1313
Rosgen, D. A. (1994). A classification of natural rivers. Catena, 22,169–199. https://doi.org/10.1016/0341-8162(94)90001-9
Rotenberry, J. T. (1985). The role of habitat in avian community composition: physiognomy or floristics? Oecologia, 67,213– 217. https://doi.org/10.1007/BF00384286
Sánchez-Montoya, M. M., Moleón, M., Sánchez-Zapata, J.A., & Escoriza, D. (2017). The biota of intermittent and ephemeral rivers: amphibians, reptiles, birds, and mammals. In T. Datry, N. Bonada, & A. J. Boulton (Eds.), Intermittent rivers and ephemeral streams: ecology and management (pp. 299–322). San Diego, CA: Academic Press.
Şekercioǧlu, C. H., Loarie, S. R., Oviedo-Brenes, F., Mendenhall, C. D., Daily, G. C., & Ehrlich, P. R. (2015). Tropical countryside riparian corridors provide critical habitat and connectivity for seed-dispersing forest birds in a fragmented landscape. Journal of Ornithology, 156,343–353. https://doi.org/10.1007/s10336-015-1299-x
Semarnat (Secretaría del Medio Ambiente y Recursos Naturales). (2010). Norma Oficial Mexicana NOM-059-SEMARNAT-2010, Protección ambiental – Especies nativas de México de flora y fauna silvestres – Categorías de riesgo y especificaciones para su inclusión, exclusión o cambio – Lista de especies en riesgo. Diario Oficial de la Federación. 30 de diciembre de 2010, Segunda Sección, México.
Seymour, C. L., & Simmons, R. E. (2008). Can severely fragmented patches of riparian vegetation still be important for arid-land bird diversity? Journal of Arid Environments, 72,2275–2281. https://doi.org/10.1016/j.jaridenv.2008.07.014
Skagen, S. K., Melcher, C. P., Howe, W. H., & Knopf, F. L. (1998). Comparative use of riparian corridors and oases by migrating birds in southeast Arizona. Conservation Biology, 12,896–909. https://doi.org/10.1111/j.1523-1739.1998.96384.x
Strong, T. R., & Bock, C. E. (1990). Bird species distribution patterns in riparian habitats in southeastern Arizona. Condor, 92,866–885. https://doi.org/10.2307/1368723
Szaro, R. C., & Jakle, M. D. (1985). Avian use of a desert riparian island and its adjacent scrub habitat. Condor, 87,511–519. https://doi.org/10.2307/1367948
Wiens, J. A., & Rotenberry, J. T. (1981). Habitat associations and community structure of birds in shrubsteppe environments. Ecological Monographs, 51,21–41. https://doi.org/10.2307/
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Perfil de leucocitos como biomarcador hematológicoen poblaciones de la salamandra de arroyo Ambystoma ordinarium
Carolina González-Pardo a, Ireri Suazo-Ortuño a, *, Cinthya Mendoza-Almeralla b, David Tafolla-Venegas c, Yurixhi Maldonado-López a y Esperanza Meléndez-Herrera a
a Universidad Michoacana de San Nicolás de Hidalgo, Instituto de Investigaciones sobre los Recursos Naturales, Avenida San Juanito Itzícuaro s/n, Nueva Esperanza, 58330 Morelia, Michoacán, México
b Universidad Autonóma del Estado de Hidalgo, Instituto de Ciencias Básicas e Ingeniería, Centro de Investigaciones Biológicas, Laboratorio de Ecología de Poblaciones, Km 4.5 carretera Pachuca-Tulancingo, 42184 Mineral de La Reforma, Hidalgo, México
c Universidad Michoacana de San Nicolás de Hidalgo, Facultad de Biología, Edificio R, Ciudad Universitaria, 58030 Morelia, Michoacán, México
*Autor de correspondencia: ireri.suazo@umich.mx (I. Suazo-Ortuño)
Recibido: 7 agosto 2023; aceptado: 22 febrero 2024
Resumen
La evaluación del perfil de leucocitos como biomarcador hematológico en las poblaciones de anfibios es cada vez más común en estudios ecológicos en especies amenazadas o en declive. En este estudio evaluamos y comparamos el perfil de leucocitos y el índice neutrófilos/linfocitos (N/L) en frotis de sangre periférica de Ambystoma ordinarium en 3 tipos de hábitats: conservados, urbanizados y agrícolas. Consideramos al perfil leucocitario como un endpoint inmunológico, ya que nos puede proporcionar información sobre la respuesta inmunológica del organismo. De acuerdo con los resultados encontrados, en los individuos de A. ordinarium de los sitios urbanizados y agrícolas se detectaron aumentos en las proporciones de eosinófilos, basófilos y monocitos, y una disminución en las proporciones de linfocitos. Asimismo, en los individuos de los sitios urbanizados y agrícolas se detectaron aumentos en el número de neutrófilos banda, además se reporta por primera vez el hallazgo de células plasmáticas en la sangre de esta especie. En general, los perfiles de leucocitos de los individuos de A. ordinarium en los sitios urbanizados y agrícolas observados en este estudio, podrían interpretarse como respuestas fisiológicas a la perturbación ambiental.
Palabras clave: Hábitats perturbados; Respuesta inmunitaria; Índice N/L; Neutrófilos banda; Células plasmáticas; Achoque michoacano
© 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/).
Leukocyte profile as hematologic biomarker in populations of the mountain salamander, Ambystoma ordinarium
Abstract
Assessing the leukocyte profile as a hematological biomarker is now frequently used in ecological studies of threatened or declining species. In this study, we evaluated and compared leukocytes profile and neutrophils/lymphocytes (N/L) ratio in peripheral blood smears of the salamander Ambystoma ordinarium in 3 types of habitats: urbanized, agricultural, and conserved. We considered leukocyte profiles as an immunological endpoint, since it can provide information about the immunological response. Results indicated that A. ordinarium individuals from the urbanized and agricultural sites presented higher proportions of neutrophils, eosinophils, basophils and monocytes and a decrease in the proportions of lymphocytes. Agricultural habitats presented higher N/L ratios. Likewise, in the individuals of urbanized and agricultural sites an increase was registered in the number of neutrophils with a band nucleus, in addition, the finding of plasma cells in the blood of this species is reported for the first time. In general, leukocyte profiles of A. ordinarium individuals in urbanized and agricultural sites observed in this study suggest that these profiles can be interpreted as physiological responses to environmental disturbance.
Keywords: Disturbed habitats; Immune response; N/L ratio; Band neutrophils; Plasma cells; Achoque michoacano
Introducción
En la actualidad, una de las preocupaciones más importantes en la conservación de vida silvestre es la pérdida y disminución global de las especies de anfibios (Alvarado, 2021). Los cambios en los hábitats asociados a las actividades antropogénicas y las enfermedades infecciosas representan las principales amenazas (Wake y Vredenburg, 2008). Sin embargo, científicos en todo el mundo, consideran que no existe una sola causa potencial, sino que éstos y otros factores pueden actuar mediante sinergias contribuyendo en la disminución de sus poblaciones (Lips et al., 2005; Stuart et al., 2004).
Los anfibios, pueden ser más vulnerables a los cambios en sus hábitats en comparación con el resto de los vertebrados, por 2 razones principales: poseen una piel delgada y porosa que es permeable al agua, y son organismos ectotermos, por lo que dependen de su entorno para conservar su temperatura (Duellman y Trueb, 1994). Por estos motivos, la evaluación del estado de salud de las poblaciones de anfibios es cada vez más común en estudios ecológicos en especies amenazadas o en declive (Barriga-Vallejo et al., 2015; Das y Mahapatra, 2014; Shutler y Marcogliese, 2011). Los perfiles de leucocitos han sido evaluados con mayor frecuencia porque proporcionan información sobre el estado inmunológico y permiten detectar cambios fisiológicos y patológicos tempranos en los individuos, sobretodo, estudios recientes han comenzado a incorporarlos como biomarcadores para evaluar la salud de los individuos y de su ambiente (Barni et al., 2007; Cabagna et al., 2005; Davis et al., 2010; Salinas et al., 2015, 2019).
Los leucocitos (linfocitos, neutrófilos, eosinófilos, basófilos y monocitos) son células sanguíneas que forman parte del sistema inmunitario, desempeñando funciones cruciales en la defensa contra infecciones y enfermedades (Thrall, 2004). De esta forma, los leucocitos en sangre pueden aumentar rápidamente en una infección, por ejemplo, aumentos en las proporciones de eosinófilos se han asociado con infecciones parasitarias (Davis y Golladay, 2019; Ramírez-Hernández et al., 2019) y en evidencia reciente, se han reportado aumentos en las frecuencias de linfocitos maduros e inmaduros (Salinas et al., 2019). El perfil de leucocitos también ha sido evaluado con éxito como indicador de estrés en poblaciones en ambientes perturbados y alteraciones morfológicas como el aumento de neutrófilos sin segmentación nuclear se han relacionado con ambientes contaminados con desechos urbanos y agrícolas (Barni et al., 2007; Cabagna et al., 2005; Ramírez-Hernández et al., 2019; Romanova y Romanova, 2003).
En México, habitan 14 especies del género Ambystoma (Ramírez-Bautista et al., 2023) y se ha estudiado el perfil de leucocitos en algunas especies como biomarcador de inflamación y estrés asociado a perturbaciones antropogénicas (Barriga-Vallejo et al., 2015; Ramírez-Hernández et al., 2019). La salamandra de arroyo Ambystoma ordinarium se distribuye en el noreste de Michoacán y se encuentra catalogada como en peligro de extinción por la IUCN (2024), y como especie protegida por el gobierno de México (Semarnat, 2010). Particularmente, en varios sitios del área de distribución de esta especie existe un fuerte impacto sobre los arroyos que habita debido a presiones de urbanización, actividades agrícolas y ganaderas (Soto-Rojas, 2012). Considerando el contexto en el que se encuentra esta especie, es importante monitorear sus poblaciones, sobre todo las que están sujetas a la continua perturbación de sus hábitats. Por lo tanto, el objetivo de este estudio fue evaluar y comparar los perfiles de leucocitos como biomarcador hematológico en poblaciones de A. ordinarium de hábitats con diferentes grados de perturbación: conservados, urbanizados y agrícolas.
Materiales y métodos
Se realizaron 2 visitas, la primera en noviembre de 2020 y la segunda en marzo de 2021 a 9 sitios con arroyos habitados por A. ordinarium condiferentes grados de perturbación en Michoacán (fig. 1).Los sitios 1, 2 y 3 se encuentran en zonas conservadas en los municipios de Charo, Morelia y Zinápecuaro. En estos sitios, la vegetación está representada por bosque de pino y pino-encino y no se encuentran afectados por la urbanización, cultivos ni zonas de pastoreo.
Los sitios 4, 5 y 6 son urbanizados y se localizan en la Ciudad de Morelia, la cual tiene una extensión de 1,333 km2 y presenta más de 500,000 habitantes (Magaña-Martínez y Reyes Camacho, 2012; tabla 1, fig. 1). Cada sitio se encuentra a una distancia mínima de 3 km uno con respecto a otro, por lo que corresponden a 3 poblaciones independientes de acuerdo a la poca vagilidad de la especie reportada por Montes-Calderón et al. (2011). Se consideraron los sitios como perturbados debido a: 1) la presencia de construcciones urbanas (López- Granados et al., 2008; Magaña-Martínez y Reyes Camacho, 2012), 2) el vertimiento de aguas contaminadas con fertilizantes y pesticidas (López- Granados et al., 2008) y 3) se encuentran a menos de 1 km de avenidas principales de la ciudad de Morelia (Téllez-Ramírez, 2012).
Los sitios 7, 8 y 9 se encuentran en zonas agrícolas en los municipios de Queréndaro, Indaparapeo y Zinapécuaro (fig. 1). Estos sitios presentan arroyos permanentes con escasa vegetación ribereña y se consideran perturbados porque durante el muestreo de este estudio, estaban rodeados de cultivos de maíz y potreros, y no presentaban vegetación de bosque de pino y pino-encino.
En cada arroyo se realizó una búsqueda intensiva de los ejemplares mediante la técnica de inspección por encuentro visual (VES) (Crump y Scott, 1994). Una vez localizados, se capturaron con red de mano y se colocaron en recipientes con agua de su medio para evitar su desecación. Inmediatamente, en el sitio de colecta, se manipuló a cada ejemplar con guantes estériles y se obtuvo una gota de sangre periférica de un pequeño corte de una de las branquias, llevando a cabo el procedimiento sin sacrificar a los individuos. La gota de sangre se colocó en el extremo de un portaobjetos limpio y con ayuda de un segundo, el cual se colocó en un ángulo de 45° por delante de la gota, se lo hizo retroceder hasta tocar la gota, luego se deslizó ejerciendo una presión suave y firme hacia delante. Cada frotis sanguíneo se secó a temperatura ambiente por 3 min y se fijó con metanol. Al final del procedimiento todos los organismos se liberaron en sus respectivos arroyos de origen. El manejo de las salamandras y las muestras se realizó con el permiso de colecta científica número SGPA/DGVS/13339/19 otorgado por la Semarnat.
Los frotis sanguíneos fueron llevados al laboratorio de Parasitología de la Universidad Michoacana de San Nicolás de Hidalgo (UMSNH) y se cubrieron con el colorante no diluido de Wright, dejándose reposar durante 5 min. Posteriormente, se les agregó solución buffer de Wright gota a gota, hasta que apareció una película metálica sobre la muestra, después se dejaron reposar por 5 min. Finalmente, los frotis fueron lavados con agua destilada, hasta que el colorante se lavó y se dejaron secar.
Figura 1. Mapa de la ubicación de los sitios de estudio de A. ordinarium en algunos municipios de Michoacán. 1. 5.7 km al este de Jaripeo, 2. Agua Zarca, 3. 8.9 km al oeste de Bocaneo, 4. Puente campestre, 5. Filtros viejos, 6. Río Chiquito, 7. 14 km al sur de Queréndaro, 8. 0.75 km al sur de Ziróndaro, 9. 10.86 km al sureste del Municipio de Queréndaro.
Tabla 1
Datos de colecta de Ambystoma ordinarium. Se muestra el nombre, las coordenadas y la categoría de los sitios de colecta y el número y talla de los organismos colectados.
Número y nombre de sitio | Coordenadas | Categoría del sitio | Individuos colectados en invierno 2020 | Individuos colectados en primavera 2021 | Longitud LHC en mm |
1. 5.7 km al este de Jaripeo | 19°40’28.3” N, 101°01’44.9” O | Conservado | 4 | 3 | 74 a 77 |
2. Agua Zarca | 19°34’28.9” N, 101°07’28.2” O | Conservado | 3 | 13 | 66 a 77 |
3. 8.9 km al oeste de Bocaneo | 19°50’28.6” N, 100°43’55.1” O | Conservado | 0 | 13 | 67 a 88 |
4. Puente campestre | 19°40’31.3” N, 101°09’27.5′′ O | Urbanizado | 3 | 3 | 79 a 98 |
5. Filtros Viejos | 19°40’01.0” N, 101°08’36” O | Urbanizado | 5 | 8 | 77 a 91 |
6. Río Chiquito | 19°36’38.2” N, 101°07’26.8” O | Urbanizado | 7 | 5 | 81 a 101 |
7. 14 km al sur de Queréndaro | 19°41’05.2” N, 100°52’31.2” O | Agricola | 5 | 11 | 75 a 94 |
8. 0.75 km al sur de Ziróndaro | 19°42’58.4” N, 100°54’59.6” O | Agricola | 9 | 8 | 67 a 118 |
9. 10.86 km al sureste del Municipio de Queréndaro | 19°45’25.4” N, 100°50’33.9” O | Agricola | 4 | 5 | 82 a 111 |
Cada frotis se observó al microscopio óptico con el aumento 400x y se efectuó el recuento diferencial de leucocitos en movimiento zigzag. Las células fueron contadas por una sola persona para evitar la variabilidad en las observaciones considerando las características morfológicas descritas por Thrall (2004), Hadji-Azimi et al. (1987) y Salinas et al. (2017). Siguiendo a Davis et al. (2008), en cada frotis sanguíneo se contaron 100 células, determinándose las proporciones relativas de los 5 tipos de leucocitos y el índice N/L propuesto como indicador de respuesta al estrés. Para la evaluación morfológica de los leucocitos, durante el recuento diferencial, se contaron los leucocitos con cambios en la coloración y presencia de manifestaciones morfológicas en el citoplasma; además, se evaluó la segmentación nuclear de los neutrófilos (neutrófilos banda, identificados por la falta de segmentación en el núcleo).
Para comparar las proporciones y la morfología de cada tipo de leucocito y los índices N/L de los individuos de A. ordinarium entre los sitios conservados, urbanizados y agrícolas, se utilizaron modelos lineales generalizados mixtos (GLM), con error de distribución Poisson debido a que las variables de respuesta son conteos. Los análisis estadísticos se realizaron en R versión 4.2.0 (R Core Team, 2022) y se usó el paquete ggplot2 versión 3.4.1 (Wickham et al., 2016) para la realización de figuras.
Resultados
Se recolectaron 109 individuos de Ambystoma ordinarium en los 3 sitios, 36 en conservados, 31 en urbanizados y 42 en los agrícolas (tabla 1), 40 ejemplares se obtuvieron en invierno de 2020 y 69 en primavera de 2021. El promedio de la longitud total de los ejemplares fue de 82.44 mm (mínima 66-118 máxima) y de acuerdo con las tallas reportadas por Anderson y Worthington (1971), todos los individuos recolectados fueron adultos metamórficos (tabla 1).
Los promedios en las proporciones de los 5 tipos de leucocitos y el índice N/L de los individuos de A. ordinarium para los sitios conservados, urbanizados y agrícolas se presentan en la tabla 2. Las proporciones de linfocitos mostraron diferencias significativas entre los sitios. Se detectaron disminuciones en las proporciones de estas células en individuos de los sitios urbanizados y agrícolas en comparación con los individuos de los hábitats conservados (fig. 2). No se detectaron diferencias en las proporciones de neutrófilos y los índices N/L entre los sitios (tabla 2). Por último, las proporciones de eosinófilos, basófilos y monocitos mostraron diferencias significativas entre los sitios (tabla 2). Se detectaron aumentos en las proporciones de estas células en individuos de los sitios urbanizados y agrícolas con respecto a los individuos de los sitios conservados (fig. 2).
Figura 2. Gráfica de cajas y alambres que muestra las diferencias en las proporciones de leucocitos entre los hábitats urbanizados, agrícolas y conservados. Las letras representan las diferencias de medias entre grupos de acuerdo al GLM.
Los linfocitos, fueron los leucocitos de menor tamaño, la mayoría de éstos se caracterizaron por ser células redondas con un núcleo central que ocupó la mayor parte del citoplasma basófilo (fig. 3). En los frotis sanguíneos de 2 individuos de los hábitas agrícolas se observaron células plasmáticas, caracterizadas por un núcleo excéntrico, citoplasma abundante con aumento en la basofília y abundantes inclusiones globulares y hialinas (fig. 3).
Tabla 2
Promedios relativos de los leucocitos e índices N/L (±error estándar) en individuos adultos de Ambystoma ordinarium entre hábitats conservados, urbanizados y agrícolas.
Variable de respuesta | Hábitats conservados | Hábitats urbanizados | Hábitats agrícolas | gl | c2 | p |
Linfocitos | 83.1 (± 0.51) | 72.8 (± 1.20) | 74.5 (± 0.93) | 2 | 27.84 | < 0.001 |
Neutrófilos | 8.4 (± 0.69) | 8.9 (± 0.95) | 9.6 (± 0.97) | 2 | 2.84 | 0.095 |
Eosinófilos | 3.9 (± 0.38) | 8.3 (± 1.53) | 9.0 (± 0.84) | 2 | 88.37 | < 0.001 |
Basófilos | 2.7 (± 0.42) | 4.9 (± 0.68) | 4.5 (± 0.52) | 2 | 26.66 | < 0.001 |
Monocitos | 1.7 (± 0.43) | 4.9 (± 1.24) | 2.2 (± 0.39) | 2 | 63.26 | < 0.001 |
Neutrófilos banda | 1.5 (± 0.24) | 3.1 (± 0.60) | 3.8 (± 0.48) | 2 | 42.27 | <0.001 |
Índice N/L | 0.10 (± 0.00) | 0.10 (± 0.011) | 0.13 (± 0.01) | 2 | 0.17 | 0.698 |
Los neutrófilos se caracterizaron por ser células redondas irregulares que pueden o no presentar finos gránulos en el citoplasma, su núcleo violeta puede no presentar segmentaciones (neutrófilos banda, fig. 3) ó ser segmentados de 2 a 5 lóbulos (fig. 3). En relación con la comparación de la segmentación nuclear de los neutrófilos entre los sitios, se encontraron variaciones significativas, detectándose incrementos en las proporciones de neutrófilos banda en individuos de los sitios urbanizados y agrícolas con respecto a los individuos de los sitios conservados (fig. 3).
Los eosinófilos fueron células con abundantes gránulos rosados cubriendo el citoplasma, presentan un núcleo violeta generalmente bilobulado (fig. 3) y en pocas ocasiones se observó unilobulado y trilobulado. Con respecto de la morfología de los basófilos, éstos se caracterizaron por ser células redondas u ovaladas con abundantes gránulos púrpuras en el citoplasma que, por lo general, cubren el núcleo redondo o bilobulado (fig. 3). Por último, los monocitos fueron los leucocitos de mayor tamaño, son células redondas con núcleo en forma de riñón o herradura y citoplasma abundante (fig. 3). No se observaron variaciones morfológicas en eosinófilos, basófilos y monocitos.
Discusión
Se evaluaron los perfiles de leucocitos en frotis de sangre de 109 individuos adultos de la salamandra Ambystoma ordinarium, de los cuales 37 se recolectaron en hábitats conservados, 30 en hábitats urbanizados y 42 en hábitats agrícolas. La morfología y coloración de los 5 tipos de leucocitos concuerdan con lo reportado para otras especies de anuros y caudados (Cabagna et al., 2005; Hadji-Azimi et al., 1987; Salinas et al., 2017). Los linfocitos fueron los leucocitos más abundantes, detectándose disminuciones en sus proporciones en individuos de A. ordinarium de los sitios urbanizados y agrícolas considerados como perturbados. En estudios recientes, la disminución de los linfocitos en sangre se ha documentado en especies de anfibios como respuesta a factores estresantes como infecciones y hábitats contaminados con pesticidas, debido a que el aumento en las hormonas del estrés (glucocorticoides) puede inducir la salida de estas células de la sangre a los tejidos linfoides (en anfibios, hígado y bazo) (Davis et al., 2008; Waye et al., 2019). En contraste, los neutrófilos por ser las principales células encargadas del ataque a agentes infecciosos son estimulados a proliferar para migrar a los sitios de inflamación. La disminución en las proporciones de los linfocitos se ha relacionado con el aumento en las proporciones de los neutrófilos (índice N/L) como biomarcador de estrés (Davis et al., 2010). En este estudio, no se detectaron variaciones significativas en las proporciones de neutrófilos y en los índices N/L promedio. En estudios previos en poblaciones silvestres de salamandras del género Ambystoma y otras especies de anfibios, se han reportado índices N/L promedio cercanos a 0.40 (Cabagna et al., 2005; Davis y Durso, 2009; Shutler et al., 2009). Sin embargo, a diferencia de nuestro estudio, Ramírez-Hernández et al. (2019) reportaron para esta especie un índice N/L promedio de 1.5. y 0.9 en hábitats perturbados y conservados, respectivamente. Las diferencias entre ambos resultados podrían deberse a que el tamaño de muestra utilizado por Ramírez-Hernández et al. (2019) fue pequeño en comparación con el tamaño de muestra utilizado en nuestro estudio. Además, nuestros resultados pueden indicar que otros factores no contabilizados están influyendo en las respuestas de estas células en los individuos muestreados. Un factor que aumenta el índice linfocitos y neutrófilos en anfibios es la infección por Batrachochytrium dendrobatidis (Bd; Davis et al., 2010; Savage et al., 2016). Recientemente, se reportó la presencia del hongo quitridio en las mismas poblaciones de A. ordinarium analizadas en este estudio (Mendoza-Almeralla et al., 2023). Los niveles de infección por Bd reportados fueron de entre 112 a 1,856 equivalentes zoosporas genómicas (EZG´s) en 2 sitios conservados; mientras que en los sitios perturbados, el grado de infección fue de 21 a 4,338 EZG´s, ésto sugiere que hay mayor grado de infección en lugares perturbados. Por lo que es importante saber si el número de neutrófilos y linfocitos es afectado por la infección del quitridio.
Figura 3. Leucocitos en frotis de sangre periférica de A. ordinarium. A) Linfocito, B) neutrófilo segmentado, C) neutrófilo banda, D) eosinófilo, E) basófilo, D) monocito, vista a 100X, G) célula plasmática con núcleo excéntrico y citoplasma abundante, vista a 40X.
Por otro lado, se detectaron incrementos de neutrófilos banda en individuos de los sitios perturbados con respecto a los sitios conservados. En procesos inflamatorios, los neutrófilos son impulsados a proliferar y para compensar esta demanda, la médula ósea libera en la sangre células inmaduras (neutrófilos banda) (Davis y Golladay, 2019). En anfibios existe evidencia de incrementos en las proporciones de estas células como respuesta inflamatoria contra fertilizantes y pesticidas (Mann et al., 2009; Romanova et al., 2022). Pese a que en este estudio no evaluamos la presencia de estos contaminantes químicos en el agua, en los sitios urbanizados seleccionados aquí, se han reportado descargas de aguas contaminadas con fertilizantes y pesticidas (López-Granados et al., 2008), y en los sitios agrícolas es probable que el uso de estos agroquímicos sea habitual y su dispersión fuera de estas áreas pueda llegar hasta los arroyos donde habita esta especie. Los incrementos detectados en las proporciones de eosinófilos, basófilos y monocitos en los individuos de los sitios urbanizados y agrícolas también se pueden asociar con la presencia de estos contaminantes químicos en sus hábitats. En algunos estudios en anfibios se han reportado incrementos en las proporciones de eosinófilos en sitios contaminados por pesticidas por su capacidad de reaccionar a antígenos ambientales (Attademo et al., 2014; Romanova y Romanova, 2003). Con respecto a los monocitos, el aumento de estas células podría relacionarse con el incremento de la fagocitosis de desechos tisulares, puesto que, de acuerdo con algunos autores la exposición prolongada a contaminantes químicos aumenta la necrosis tisular (Zhelev, 2007). El papel de los basófilos en las respuestas inmunitarias de los anfibios no es claro (Allender y Fry, 2008), sin embargo, al igual que en otros grupos de vertebrados parecen desempeñar un papel importante en la inflamación (Claver y Quaglia, 2009).
Los estudios realizados por varios investigadores han revelado que los anfibios son capaces de generar una respuesta inmunitaria a antígenos complejos asociados con patógenos y antígenos ambientales (Grogan et al., 2018; Savage y Zamudio, 2011; Zhelev, 2007). Los linfocitos (células T y B), ante la presencia de antígenos son responsables de activar la inmunidad mediada por células específicas de patógenos (células T citotóxicas o auxiliares) y la inmunidad humoral (células plasmáticas secretoras de anticuerpos específicos) (Grogan et al., 2018). El hallazgo de células plasmáticas en la sangre de los individuos de los sitios agrícolas podría atribuirse a la especificidad de la respuesta inmune humoral a antígenos tóxicos presentes en el ambiente. En un estudio previo, Zhelev (2007) reportó aumentos de estas células en individuos de Rana ridibunda en hábitats con contaminación industrial. El hallazgo de estas células es interesante, porque casi no hay datos en la literatura sobre su presencia y apariencia en la sangre de los anfibios. Aunque no podemos asegurar que se trate de este tipo de células, estos datos sin duda sientan las bases para desarrollar estudios bioquímicos e inmunológicos futuros.
En anfibios se ha documentado la función antipara-
sitaria de los eosinófilos en infecciones con nemátodos y tremátodos (Belden y Kiesecker, 2005; Davis y Golladay, 2019; Kiesecker, 2002; Rohr et al., 2008). Particularmente, Ramírez-Hernández et al. (2019) reportaron aumentos en las cargas parasitarias en poblaciones de A. ordinarium por 2 especies de tremátodos, Gorgoderina attenuata y Ochetosoma sp.,y 2 especies de nemátodos Cosmocercoides sp.y Hedruris siredonis en hábitats perturbados, siendo uno de estos hábitats correspondiente al sitio Río Chiquito de este estudio. Adicionalmente, Mendoza-Almeralla et al. (2023) reportaron la infección por el nemátodo del género Capillaria en el sitio Filtros viejos. Por lo tanto, los incrementos de eosinófilos detectados en los individuos de estos sitios, podrían relacionarse con la presencia de mayor prevalencia de infecciones parásitarias.
Finalmente, los perfiles de leucocitos de los individuos de A. ordinarium en este estudio proporcionan parámetros hematológicos de comparación entre distintas poblaciones. De acuerdo con nuestros resultados, la evaluación de los perfiles de leucocitos es uno de los métodos más simples y menos invasivos. Los cambios en sus valores, especialmente en las poblaciones de hábitats perturbados, pueden utilizarse con éxito en evaluaciones futuras para detectar cambios fisiológicos y patológicos tempranos en los individuos y puede ser una señal de advertencia de degradación ambiental. Sin embargo, es importante reiterar que interpretar las proporciones de los leucocitos puede ser complicado, debido a que los leucocitos pueden responder a diversos factores (Barni et al., 2007; Cabagna et al., 2005; Romanova y Romanova, 2003; Shutler y Marcogliese, 2011; Shutler et al., 2009). Por ello es necesario llevar a cabo estudios complementarios sobre enfermedades infecciosas, calidad del agua, niveles de hormonas de estrés, presencia de pesticidas o metales pesados, entre otros, que permitan relacionar y estudiar la respuesta de los leucocitos en las poblaciones de esta especie y otras especies de anfibios con los diversos factores o contextos en los que se encuentran estas especies.
Agradecimientos
Este estudio fue parte del proyecto “Descifrando el microbioma de la piel en ajolotes y las consecuencias de la interacción huésped microbioma sobre una enfermedad letal emergente” de la Secretaría de Educación Pública/Consejo Nacional de Humanidades, Ciencias y Tecnologías Ciencias de Frontera. FORDECYT-PRONACES/373914/2020. Los resultados de este estudio forman parte de la tesis de maestría del autor principal, bajo la dirección de ISO y CMA. CGP agradece el apoyo financiero del Programa Nacional de Becas de SEP/Conahcyt.
Referencias
Allender, M. C. y Fry, M. M. (2008). Amphibian hematology. Veterinary Clinics of North America: Exotic Animal Practice, 11, 463–480. https://doi.org/10.1016/j.cvex.2008.03.006
Alvarado, J. G. A. (2021). Anfibios en peligro: amenazas y estrategias efectivas de conservación. Biocenosis, 32, 3–45. https://doi.org/10.22458/rb.v32i1.3552
Anderson, J. D. y Worthington, R. D. (1971). The life history of the Mexican salamander Ambystoma ordinarium Taylor. Herpetologica, 27, 165–176.
Attademo, A. M., Peltzer, P. M., Lajmanovich, R. C., Cabagna-Zenklusen, M. C., Junges, C. M. y Basso, A. (2014). Biological endpoints, enzyme activities, and blood cell parameters in two anuran tadpole species in rice agroecosystems of mid-eastern Argentina. Environmental Monitoring and Assessment, 186, 635–649. https://doi.org/10.1007/s10661-013-3404-z
Barni, S., Boncompagni, E., Grosso, A., Bertone, V., Freitas, I., Fasola, M. et al. (2007). Evaluation of Rana snk esculenta blood cell response to chemical stressors in the environment during the larval and adult phases. Aquatic Toxicology, 81, 45–54. https://doi.org/10.1016/j.aquatox.2006.10.012
Barriga-Vallejo, C., Hernández-Gallegos, O., Von-Herbing, I. H., López-Moreno, A. E., Ruiz-Gómez, M. D. L., Granados-González, G. et al. (2015). Assessing population health of the Toluca Axolotl Ambystoma rivulare (Taylor, 1940) from México using leukocyte profiles. Herpetological Conservation and Biology, 10, 592–601.
Belden, L. K. y Kiesecker, J. M. (2005). Glucocorticosteroid hormone treatment of larval treefrogs increases infection by Alaria sp. trematode cercariae. Journal of Parasitology, 91, 686–688. https://doi.org/10.1645/GE-397R.
Cabagna, M. C., Lajmanovich, R. C., Stringhini, G., Sánchez-Hernández, J. C. y Peltzer, P. M. (2005). Hematological parameters of health status in the common toad Bufo arenarum in agroecosystems of Santa Fe Province, Argentina. Applied Herpetology, 2, 373–380. https://doi.org/10.1163/157075405774483085
Claver, J. A. y Quaglia, A. I. (2009). Comparative morphology, development, and function of blood cells in nonmammalian vertebrates. Journal of Exotic Pet Medicine, 18, 7–97. https://doi.org/10.1053/j.jepm.2009.04.006
Crump, M. L. y Scott, N. Y. (1994). Visual encounter surveys. En W. Heyer, M. A. Donnelley, R. A. Mcdiarmid, L. C. Hayek. y M. C. Foster (Eds.), Measuring and monitoring biological diversity: standard methods for amphibians (pp. 84–92). Washington D.C.: Smithsonian Institution.
Das, M. y Mahapatra, P. K. (2014). Hematology of wild caught Dubois’s tree frog Polypedates teraiensis, Dubois, 1986 (Anura: Rhacophoridae). The Scientific World Journal, 491415, 7. https://doi.org/10.1155/2014/491415
Davis, A. K. y Durso, A. M. (2009). White blood cell differentials of northern cricket frogs (Acris c. crepitans) with a compilation of published values from other amphibians. Herpetologica, 65, 260–267. https://doi.org/10.1655/08-052R1.1
Davis, A. K. y Golladay, C. (2019). A survey of leukocyte profiles of red-backed salamanders from Mountain Lake, Virginia, and associations with host parasite types. Comparative Clinical Pathology, 28, 1743–1750. https://doi.org/10.1007/s00580-019-03015-9
Davis, A. K., Maney, D. L. y Maerz, J. C. (2008). The use of leukocyte profiles to measure stress in vertebrates: a review for ecologists. Functional Ecology, 22, 760–772. https://doi.org/10.1645/GE-397R
Davis, A. K., Keel, M. K., Ferreira, A. y Maerz, J. C. (2010). Effects of chytridiomycosis on circulating white blood cell distribu-
tions of bullfrog larvae (Rana catesbeiana). Comparative Clinical Pathology, 19, 49–55. 10.1007/s00580-009-0914-8
Duellman, W. E. y L. Trueb. (1994). The biology of amphibians. Baltimore, Maryland: Johns Hopkins University Press.
Grogan, L. F., Robert, J., Berger, L., Skerratt, L. F., Scheele, B. C., Castley, J. G. et al. (2018). Review of the amphibian immune response to chytridiomycosis, and future directions. Frontiers in Immunology, 9, 2536. https://doi.org/10.3389/fimmu.2018.02536
Hadji-Azimi, I., Coosemans, V. y Canicatti, C. (1987). Atlas of Xenopus laevis laevis hematology. Developmental and Comparative Immunology, 11, 807–874.
IUCN. (2020). Ambystoma ordinarium. The IUCN Red List of Threatened Species. Recuperado el 01 junio, 2024 de: https://www.iucnredlist.org/es/species/59066/161153310
Kiesecker, J. M. (2002). Synergism between trematode infection and pesticide exposure: a link to amphibian deformities in nature? Proceedings of the National Academy of Sciences, 99, 9900–9904. https://doi.org/10.1073/pnas.152098899
Lips K. R., Burrowes, P. A., Mendelson, J. R. y Parra-Olea, G. (2005). Amphibian population declines in Latin America: a synthesis. Biotropica, 37, 222–226. https://doi.org/10.1111/j.1744-7429.2005.00029.x
López-Granados, E., Mendoza-Cantú, M., Bocco, G. y Espinosa-Bravo, M. (2008). Patrones de degradacion ambiental en la Cuenca del Lago de Cuitzeo, Michoacán. Una perspectiva espacial. Centro de Investigaciones en Ecosistemas, UNAM/Morelia, Michoacán, Instituto de Geografía, UNAM/ Dirección General de Investigación de Ordenamiento Ecológico y Conservación de los Ecosistemas, INIFAP.
Magaña-Martínez, H. M. y Reyes-Camacho, E. (2012). Parque lineal ecológico los filtros viejos en Morelia, Mich. (Tesina). Facultad Arquitectura, Universidad Michoacana de San Nicolás de Hidalgo. Morelia, Michoacán.
Mann, R. M., Hyne, R. V., Choung, C. B. y Wilson, S. P. (2009). Amphibians and agricultural chemicals: review of the risks in a complex environment. Environmental Pollution, 157, 2903–2927. https://doi.org/10.1016/j.envpol.2009.05.015
Mendoza-Almeralla, C., Tafolla-Venegas, D., González-Pardo,
C. y Suazo- Ortuño, I. (2023). Primer registro de infección por Batrachochytrium dendrobatidis y por el nemátodo del género Capillaria y la ausencia de infección por Ribeiroia ondatrae en Ambystoma ordinarium. Revista Latinoamericana de Herpetología, 6, e615-05. https://doi.org/10.22201/fc.25942158e.2023.4.615
Montes-Calderón, A. M., Alvarado-Díaz, J. y Suazo-Ortuño, I. (2011). Abundancia, actividad espacial y crecimiento de Ambystoma ordinarium Taylor 1940 (Caudata: Ambys-
tomatidae) en Michoacán, México. Revista Biológicas, 13, 50–53.
Ramírez-Bautista, A., Torres-Hernández, L. A., Cruz-Elizalde, R., Berriozabal-Islas, C., Hernández-Salinas, U., Wilson, L. D. et al. (2023). An updated list of the Mexican herpetofauna: with a summary of historical and contemporary studies. Zookeys, 1166, 287. https://doi.org/10.3897/zookeys.1166.86986
Ramírez-Hernández, G., Suazo-Ortuño, I., Alvarado-Díaz, J., Escalera-Vázquez, L. H., Maldonado-López, Y. y Tafolla-Venegas, D. (2019). Effects of habitat disturbance on parasite infection and stress of the endangered Mexican stream salamander Ambystoma ordinarium. Salamandra, 55, 160–172.
R Core Team. (2022). R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. Recuperado el 01 junio, 2023 de: https://www.R-project.org/
Rohr, J. R., Schotthoefer, A. M., Raffel, T. R., Carrick, H. J., Halstead, N., Hoverman, J. T. et al. (2008). Agrochemicals in-
crease trematode infections in a declining amphibian species. Nature, 455, 1235–1239. https://doi.org/10.1038/nature07281
Romanova, E. B. y Romanova, O. Y. (2003). Peculiarities of leukocytic formula of peripheral blood of green frogs under conditions of anthropogenic load. Journal of Evolutionary Biochemistry and Physiology, 39, 480–484. https://doi.org/10.1023/B:JOEY.0000010246.27310.85
Romanova, E. B., Stolyarova, I. A., Bakiev, A. G. y Gorelov, R. A. (2022). The leukocyte blood composition of Emys orbicularis and Mauremys caspica (Reptilia: Testudines: Emydidae, Geoemydidae) at syntopy. Biology Bulletin, 49, 1923–193. https://doi.org/10.35885/1684-7318-2022-1-79-93
Salinas, Z. A., Salas, N. E., Baraquet, M. y Martino, A. L. (2015). Biomarcadores hematológicos del sapo común Bufo (Rhinella) arenarum en ecosistemas alterados de la provincia de Córdoba. Acta Toxicológica Argentina, 23, 25–35.
Salinas, Z. A., Baraquet, M., Grenat, P. R., Martino, A. L. y Salas, N. E. (2017). Morphology and size of blood cells of Rhinella arenarum (Hensel, 1867) as environmental health assessment in disturbed aquatic ecosystem from central Argentina. Environmental Science and Pollution Research, 24, 24907–24915.
Salinas, Z. A., Babini, M. S., Grenat, P. R., Biolé, F. G., Martino, A. L. y Salas, N. E. (2019). Effect of parasitism of Lernaea cyprinacea on tadpoles of the invasive species Lithobates catesbeianus. Heliyon, 5, 6. https://doi.org/10.1016/j.heliyon.
2019.e01834
Savage, A. E. y Zamudio, K. R. (2011). MHC genotypes associate with resistance to a frog-killing fungus. Proceedings of the National Academy of Sciences, 108, 16705–16710. https://doi.org/10.1073/pnas.1106893108
Savage, A. E., Terrell, K. A., Gratwicke, B., Mattheus, N. M., Augustine, L. y Fleischer, R. C. (2016). Reduced immune function predicts disease susceptibility in frogs infected with a deadly fungal pathogen. Conservation Physiology, 4, cow011. https://doi.org/10.1093/conphys/cow011
Semarnat (Secretaría del Medio Ambiente y Recursos Naturales). (2010). Norma Oficial Mexicana NOM-059-SEMARNAT-2010, Protección ambiental – Especies nativas de México de flora y fauna silvestres – Categorías de riesgo y especificaciones para su inclusión, exclusión o cambio – Lista de especies en riesgo. Diario Oficial de la Federación. 30 de diciembre de 2010, Segunda Sección, México.
Shutler, D. y Marcogliese, D. J. (2011). Leukocyte profiles of northern leopard frogs, Lithobates pipiens, exposed to pesticides and hematozoa in agricultural wetlands. Copeia, 2, 301–307. https://doi.org/10.1643/CP-10-065
Shutler, D., Smith, T. G. y Robinson, S. R. (2009). Relationships between leukocytes and Hepatozoon spp. in green frogs, Rana clamitans. Journal of Wildlife Diseases, 45, 67–72. https://doi.org/10.7589/0090-3558-45.1.67
Soto Rojas, C. (2012). Uso y selección del microhábitat de la salamandra de montaña Ambystoma ordinarium (Tesis de maestría). Universidad Michoacana de San Nicolás de Hidalgo, Morelia, Michoacán.
Stuart, S. N., Chanson, J. S., Cox, N. A., Young, B. E., Rodrigues, A. S., Fischman, D. L. et al. (2004). Status and trends of amphibian declines and extinctions worldwide. Science, 306, 1783-1786. https://doi.org/10.1126/science.1103538
Tellez-Ramirez, K. L. (2012). Programa de revitalización del Área Natural Protegida de los Filtros Viejos. Monografía para obtener el título de arquitecto. Universidad Vasco de Quiroga, Morelia, Michoacán.
Thrall, M. A. (2004). Hematology of reptiles. EnD. B. Baker, T. C. Campbell, D. DeNicola, M. J. Fettman, E. D. Lassen, A. Rebar et al. (Eds.), Veterinary hematology and Clinical Chemistry: text and clinical case presentations. Philadelphia: Lippincott Williams y Wilkins.
Wake, D. B. y Vredenburg, V. T. (2008). Are we in the midst
of the sixth mass extinction? A view from the world of amphibians. Proceedings of the National Academy of
Sciences, 105, 11466–11473. https://doi.org/10.1073/pnas.
0801921105
Waye, H. L., Dolan, P. C. y Hernández, A. (2019). White blood cell profiles in long-term captive and recently captured eastern tiger salamanders (Ambystoma tigrinum). Copeia, 107, 138–143. https://doi.org/10.1643/CP-18-126
Wickham, H., Chang, W. y Wickham, M. H. (2016). Package ‘ggplot2’. Create elegant data visualizations using the grammar of graphics, Version 2, 1–189.
Zhelev, Z. M. (2007). Investigation on the blood differential formula in Rana ridibunda (Anura, Amphibia) from the Area of the Maritsa-Iztok 1 Steam Power Plant. Acta Zoologica Bulgarica, 59, 181–190.
Estructura comunitaria de bivalvos (Mollusca: Bivalvia) asociados a macroalgas intermareales de Guerrero, México
Fernando Arriola-Álvarez a, Luis Gabriel Aguilar-Estrada a, *, Lucía Álvarez-Castillo b,
Ivette Ruiz-Boijseauneau a y Dení Rodríguez a
a Universidad Nacional Autónoma de México, Facultad de Ciencias, Laboratorio de Ficología (Biodiversidad Marina), Circuito Exterior s./n., Coyoacán, 04510 Ciudad de México, México
b Universidad Nacional Autónoma de México, Facultad de Ciencias, Posgrado en Ciencias del Mar y Limnología, Ciudad Universitaria, Coyoacán, 04510 Ciudad de México, México
*Autor para correspondencia: lgae@ciencias.unam.mx (L.G. Aguilar-Estrada)
Recibido: 7 julio 2022; aceptado: 11 septiembre 2024
Resumen
Las macroalgas intermareales proporcionan alimento y refugio para diferentes organismos. El objetivo de este trabajo fue analizar los bivalvos asociados a macroalgas. Se realizaron muestreos en enero, mayo, julio y noviembre de 2014, se recolectaron manualmente 72 muestras de macroalgas y sus bivalvos asociados dentro de cuadros de 400 cm2 en 2 localidades de Guerrero: playas El Palmar y Las Gatas. La estructura comunitaria de los bivalvos se determinó a partir de la riqueza específica, composición, abundancia, distribución e índices comunitarios: diversidad de Shannon, equidad de Pielou y dominancia de Simpson. Cada especie de macroalga (59 spp.) se asoció con la propuesta de grupos morfofuncionales. Se analizó la cobertura de macroalgas, abundancia de bivalvos y sedimento retenido. Del total de individuos (873), se reconocieron 17 especies de bivalvos. El índice de Shannon fue de 2.15 bits/individuo. Los bivalvos se asociaron a 3 grupos morfofuncionales de macroalgas. La abundancia de bivalvos y los sedimentos retenidos disminuyeron por mes, mientras que la abundancia, cobertura y sedimentos disminuyeron al aumentar el nivel de marea. Estudios como este proporcionan información importante para el conocimiento de la diversidad costera, en este caso de una zona turística en Guerrero.
Palabras clave: Malacofauna; Pacífico tropical mexicano; Riqueza específica; Sedimentos
© 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/).
Community structure of bivalves (Mollusca: Bivalvia) associated with intertidal macroalgae of Guerrero, Mexico
Abstract
Intertidal macroalgae provide food and shelter for different organisms. The objective of this work was to analyze the bivalves associated with macroalgae. Sampling was carried out in January, May, July, and November 2014, 72 samples of macroalgae and their associated bivalves were manually collected within 400 cm2 in 2 locations in Guerrero: El Palmar and Las Gatas beaches. The community structure of bivalves was determined from specific richness, composition, abundance, distribution, and community indices: Shannon diversity, Pielou evenness and Simpson dominance. Each macroalgal species (59 spp.) was associated with proposed morphofunctional groups. Macroalgal cover, bivalve abundance and retained sediment were analyzed. Of the total number of individuals (873), 17 bivalve species were recognized. The Shannon index was 2.15 bits/individual. Bivalves were associated with 3 morphofunctional groups of macroalgae. Bivalve abundance and retained sediment decreased by month, while abundance, cover, and sediment decreased with increasing tide level. Studies like this provide important information for understanding coastal diversity, in this case of a tourist area in Guerrero.
Keywords: Malacofauna; Tropical Mexican Pacific; Species richness; Sediments
Introducción
En el Pacífico tropical mexicano se han descrito 2 patrones principales de circulación de corrientes oceánicas: primavera (marzo-abril) y otoño (septiembre-octubre), que generan variaciones espacio-temporales como las temporadas de lluvias y secas o fenómenos climatológicos (Baumgartner y Christensen, 1985; Pérez, 2013; Vega et al., 2008; Wyrtki, 1966). La zona intermareal rocosa es un sitio de transición entre los ambientes terrestre y marino, la cual se encuentra sujeta a cambios constantes de las variables ambientales como la oscilación de la marea, la intensidad lumínica, el viento, las variaciones en la salinidad y la temperatura (Salazar-Vallejo y González, 1990; Vassallo et al.,2014); ésto genera un hábitat heterogéneo con diversos microambientes en donde varios grupos de organismos pueden desarrollarse (Flores-Garza etal., 2011, 2014). Los organismos más frecuentes en esta zona son las macroalgas (Lee, 2008), mismas que proporcionan refugio y alimento para numerosos grupos de invertebrados (García-Robledo et al.,2008; Jover-Capote y Diez, 2017; Moreno, 1995; Steneck y Watling, 1982; Yang et al., 2007). Las macroalgas son un ambiente espacialmente heterogéneo, lo que hace posible que puedan albergar distintos grupos de invertebrados a lo largo del tiempo (Benedetti-Cecchi et al., 2001; Olabarria y Chapman, 2001). Los anfípodos, poliquetos y moluscos son los grupos más importantes al interior de la comunidad de macroalgas, ya que representan 70% de la abundancia en éstas (Aguilera, 2011; Colman, 1940).
Dentro del phylum Mollusca, los bivalvos son la segunda clase más representativa (Gosling, 2015). En la zona intermareal pueden vivir adheridos a diversos sustratos como rocas, arena o macroalgas (García-Cubas y Reguero, 2007). Algunos organismos que conforman la clase Bivalvia son sésiles y tienen diferentes estrategias en cuanto a sus tipos de alimentación: suspensívoros o detritívoros (Coan y Valentich-Scott, 2006). Además, generan redes de mucus para atrapar las partículas que flotan en la columna de agua (Jorgensen, 1996; Ward et al., 1998), por lo que desempeñan papeles ecológicos importantes en los cuerpos de agua; por sus hábitos de vida, son un grupo de especial interés en los estudios ecológicos (Lozada, 2015; Vega et al., 2008).
En general, los trabajos sobre moluscos en México son numerosos. En el de Sánchez (2014) se mencionó que existen alrededor de 47 contribuciones tomando en cuenta las costas del Atlántico y del Pacífico. En Guerrero se cuenta con alrededor de 30 estudios malacológicos (Gama, 2019), los cuales en su mayoría se han orientado a conocer la riqueza y composición de especies (Flores, 2004; Flores-Rodríguez et al., 2012; Lesser, 1984; López-Rojas et al., 2017). Gran parte (70%) de los trabajos en este estado se han realizado en el área de Acapulco y se han analizado distintos aspectos de la comunidad de moluscos, incluyendo a los bivalvos (Barba-Marino et al., 2010; Castro-Mondragón et al.,2016; Flores-Garza et al., 2010, 2011, 2012, 2014; Flores-Rodríguez et al.,2003; Galeana-Rebolledo et al., 2012, 2018; Garcés, 2011; Kuk-Dzul et al.,2019; Torreblanca, 2010; Torreblanca-Ramírez et al., 2012; Valdés-González etal.,2004; Villegas-Maldonado et al., 2007; Villalpando, 1986).
En la parte norte de Guerrero, que incluye Ixtapa-Zihuatanejo, se han realizado estudios ecológicos y sobre ciclos reproductivos de moluscos (Baqueiro, 1979; Flores-Rodríguez et al.,2007; Salcedo-Martínez et al.,1988), o sobre especies de importancia comercial como el de Cerros-Cornelio et al. (2021), quienes mencionaron 24 especies de moluscos, de las cuales 13 son de la clase Bivalvia. Por su parte, en la costa sur de Guerrero existen 9 trabajos malacológicos; Flores-Garza et al. (2007) analizaron la densidad de Plicopurpura columellaris (Lamarck, 1816) y su malacofauna asociada, reportando 34 especies de moluscos, de las cuales 7 fueron bivalvos.
Entre los estudios realizados en las costas de Guerrero, resalta el de Salcedo-Martínez et al. (1988) por ser el primero sobre la relación entre las macroalgas e invertebrados de Zihuatanejo, donde los moluscos, en especial la clase Gastropoda, fueron el componente mayoritario (38.72%). Existen algunos trabajos sobre la asociación alga-molusco en Ixtapa-Zihuatanejo (Aguilar-Estrada et al., 2017, 2022; Cisneros, 2016; Gama-Kwick et al., 2021; Quiroz-González et al., 2020); sin embargo, dichos estudios están enfocados a otras clases de moluscos (gasterópodos y quitones); por lo que, el conocimiento sobre la relación de las macroalgas y bivalvos es escaso. La presente contribución tiene como objetivo aportar conocimiento de la estructura comunitaria de los moluscos bivalvos asociados a macroalgas, en un ciclo anual en la zona intermareal rocosa de Ixtapa-Zihuatanejo en Guerrero.
Materiales y métodos
Se realizaron 4 salidas de campo durante enero, mayo, julio y noviembre de 2014 a Ixtapa-Zihuatanejo, Guerrero, con el propósito de observar los posibles cambios de la estructura comunitaria de los bivalvos asociados a las macroalgas en la zona. Las comunidades de macroalgas se recolectaron en la zona intermareal rocosa en 2 localidades: playa El Palmar en Ixtapa (17°39’0.4” N, 101°36’2.79” O) y el pretil de playa Las Gatas, al interior de la bahía de Zihuatanejo (17°37’22.07” N, 101°33’4.85” O) (fig. 1A). La recolección de ejemplares se realizó con un permiso otorgado por la Secretaría de Agricultura, Ganadería, Desarrollo Rural, Pesca y Alimentación (SAGARPA) (Registro Nacional de Pesca y Acuacultura -DF00000208).
Playa El Palmar es un sitio expuesto, se encuentra frente al complejo turístico Ixtapa y la playa está formada por litoral rocoso y arenoso, tiene una longitud de 2.7 km. Este lugar consta de un relieve heterogéneo, compuesto por riscos y morros de diferente tamaño y forma irregular, la zona intermareal rocosa tiene una amplitud aproximada de 2 m (Aguilar, 2017). En esta playa se han descrito patrones con una circulación del agua dominante hacia el norte a lo largo de la zona, por lo que el oleaje es intenso en la parte norte (Trasviña y Andrade, 2002) (fig. 1B).
Playa Las Gatas es un sitio protegido porque se localiza al interior de la bahía de Zihuatanejo; tiene una extensión de 350 m y se compone principalmente de arenas, fragmentos de coral y rocas (Cisneros, 2016; López, 1993). Paralelo a la línea de costa se encuentra el “pretil”, que es un conglomerado de rocas apiladas, irregulares y de tamaños variables (Aguilar, 2017; Cisneros, 2016; Urbano, 2004). La zona intermareal tiene una amplitud aproximada de 1 m (López, 1993). La presencia de este conglomerado hace que el oleaje sea menor al interior de la playa (Aguilar, 2017) (fig. 1C).
Figura 1. A, Ubicación de la zona de estudio en Ixtapa-Zihuatanejo; B, playa El Palmar; C, playa Las Gatas.
En cada localidad, 3 muestras de macroalgas se recolectaron manualmente de forma aleatoria con una espátula (Bakus, 2007), dentro de cuadros de 20 × 20 cm (0.04 m2) por cada nivel de marea: intermareal bajo (3), medio (3) y alto (3); en total, se obtuvieron 72 muestras provenientes de 36 cuadros en playa Las Gatas y de 36 cuadros en playa El Palmar. Las muestras de las comunidades de macroalgas y moluscos asociados se preservaron en una mezcla de formaldehído al 4% con agua de mar, neutralizada con borato de sodio y glicerina; después se trasladaron al Laboratorio de Ficología (Biodiversidad Marina) de la Facultad de Ciencias de la Universidad Nacional Autónoma de México (UNAM).
En el laboratorio, se calculó la cobertura de cada especie de macroalga en cm2 al colocar cada muestra sobre un área delimitada de 20 × 20 cm. Para su identificación taxonómica se tomaron en cuenta las características morfológicas externas, como tipo de talo y tipo de ramificación, e internas a partir de cortes anatómicos transversales de ejes, frondas y, de estar presentes, de estructuras reproductivas; los especímenes y cortes fueron observados bajo microscopios estereoscópico y óptico (Zeiss). La identificación taxonómica de los ejemplares de macroalgas se realizó utilizando literatura especializada para macroalgas del océano Pacífico (Abbott, 1999; Abbott y Hollenberg, 1976; Dawson, 1949, 1953, 1954, 1960, 1961, 1963; Dawson y Beaudette, 1959; Rodríguez et al.,2008; Taylor, 1945). La actualización de la nomenclatura se hizo a partir de la base de datos de Algaebase (Guiry y Guiry, 2024) y, con base en ella, se elaboró una lista sistemática de las especies de macroalgas. De cada comunidad de macroalgas se separó el sedimento retenido por las algas después de la medición de la cobertura para cada especie y se midió su peso húmedo con una balanza digital, modelo OBI.
De cada muestra, se extrajeron de forma manual todos los moluscos de la clase Bivalvia. Los ejemplares con concha y parte blanda (vivos) fueron identificados al nivel taxonómico más bajo posible, género o especie, dependiendo del estado de conservación de cada ejemplar, a partir de la observación de las características morfológicas de la concha, con apoyo de un microscopio estereoscópico. La identificación taxonómica de las especies se hizo con literatura malacológica especializada para la zona del océano Pacífico oriental (Coan et al., 2000; Keen, 1971). Se elaboró una lista sistemática con base en la propuesta de Bouchet et al. (2010) para los niveles suprafamiliares y la actualización de nomenclatura se realizó a partir de la base de datos de World Register of Marine Species (WORMS) para género y especie (Horton et al., 2024).
Los ejemplares de moluscos y macroalgas recolectados fueron depositados en la colección “Invertebrados asociados a macroalgas”, en proceso de registro, con número de inventario para bivalvos (INV-1531 a INV-1638) del Laboratorio de Ficología (Biodiversidad Marina) de la Facultad de Ciencias, UNAM y las macroalgas se depositaron en la colección del Herbario de la Facultad de Ciencias (FCME) con número de catálogo (PTM-10534 a PTM-10558; PTM-10577 a PTM-10604; PTM-10613 a PTM-10630; PTM-10640 a PTM-10648).
Se elaboró una curva de acumulación de las especies de bivalvos asociados a macroalgas recolectadas en las localidades de estudio, con la finalidad de conocer la cantidad de especies que faltaría encontrar y recolectar en dichas localidades. Con los datos de riqueza de especies (S) y abundancia de bivalvos (N) de ejemplares vivos (concha y parte blanda), se estimaron los índices de Shannon (H´) y de diversidad máxima (H’ max) para cada fecha de muestreo, ya que éstos corresponden a los miembros de la comunidad en el momento de muestreo (Aguilar-Estrada et al., 2014); H´ fue expresado en bits/individuo (Magurran, 2004). Se calculó el índice de equidad de Pielou (J´) y el índice de dominancia de Simpson (D) (Moreno, 2001). Estos índices permiten hacer comparaciones cuantitativas y cualitativas entre estudios o zonas ya que se han utilizado como referente mínimo para describir la estructura comunitaria de un lugar (Aguilar-Estrada et al., 2014).
Para evaluar la normalidad de los datos de abundancia de bivalvos y de los índices de diversidad se realizaron pruebas de Shapiro-Wilk (W) (Siegel, 1990). Posteriormente, se realizaron pruebas de Levene (F) para comprobar la homogeneidad de varianzas para los datos no normales y prueba de Bartlett para los datos con distribución normal (Bartlett, 1937; Levene, 1960). Los índices de diversidad de Shannon fueron analizados usando una prueba de “t de student” para evaluar si existían diferencias estadísticamente significativas entre los meses de muestreo. Estos análisis se realizaron utilizando la paquetería Vegan Versión 2.5-6 (Oksanen et al., 2019) en el programa R Studio 2023.12.0+369 (R Core Team, 2023).
Los datos de abundancia de bivalvos no fueron normales de acuerdo con las pruebas de normalidad, por ello, se realizaron análisis de estadística no paramétrica en el software PRIMER v6 + add on Permanova (PRIMER-E Ltd., Plymouth, UK) (Anderson et al., 2008; Clarke y Gorley, 2006). Los datos fueron transformados con 4√ y con ellos se calculó una matriz de similitud a partir del índice de Bray-Curtis. A partir de dicha matriz, se realizó un análisis de escalamiento multidimensional no métrico (nMDS) con la finalidad de observar la distribución de las abundancias de bivalvos en el área de estudio. Se incluyó la abundancia de las especies dominantes en forma de vectores azules superpuestos en el gráfico nMDS para facilitar la interpretación de las abundancias de especies dominantes en las distintas localidades y niveles de marea, donde el círculo azul representa la variación de la abundancia. Las especies dominantes que se seleccionaron en este análisis son las que estuvieron presentes en todas las muestras con abundancias mayores a la media.
Además, se realizaron análisis de varianza permutacionales (Permanova) para determinar si existían diferencias significativas en la abundancia de bivalvos (variable dependiente) con respecto a los factores utilizados: localidad con 2 niveles (El Palmar y Las Gatas), nivel de marea con 3 niveles (alto, medio y bajo) y mes de muestreo con 4 niveles (enero, mayo, julio y noviembre). En los análisis de Permanova se utilizaron 999 permutaciones de los residuos bajo un modelo reducido. Posteriormente, para los factores donde se obtuvieron diferencias significativas, se realizaron comparaciones de pares para identificar los niveles que eran diferentes estadísticamente (Anderson et al., 2008).
Utilizando las variables numéricas, se realizaron regresiones lineales múltiples con el software SPSS Statistics v20, para evaluar el efecto de la cobertura de macroalgas y los sedimentos retenidos (variables independientes) en la abundancia de bivalvos (variable dependiente) en las playas El Palmar y Las Gatas, en los diferentes niveles de marea y en los meses de muestreo.
Por último, se determinó el grupo morfofuncional (GMF): filamentosas (Fil), foliosas (Fol), foliosas corticadas (Foc), filamentosas corticadas (Fic), coriáceas (Cor), calcáreas articuladas (Cal) para cada especie de macroalgas, con base en la propuesta de Steneck y Dethier (1994) y se asociaron con las especies de bivalvos recolectadas.
Resultados
Riqueza y composición de moluscos. En 60 de las 72 muestras recolectadas se encontraron bivalvos, 35 muestras pertenecientes a playa El Palmar y 25 a playa Las Gatas. Se obtuvieron un total de 873 individuos de la clase Bivalvia. Se identificaron 17 especies (fig. 2) agrupadas en 2 subclases, 8 órdenes, 10 familias y 15 géneros. Del total de especies recolectadas, 2 de ellas Parvilucina approximata (Dall, 1901) y Pinna rugosa G. B. Sowerby I, 1835, encontradas en playa El Palmar y playa Las Gatas, respectivamente, no se incluyeron en los análisis de estructura comunitaria, dado que solo se recolectó la concha. Se obtuvieron 13 especies en playa El Palmar y 9 en playa Las Gatas. Las familias Carditidae y Mytilidae fueron las mejor representadas con 3 (18%) y 5 (30%) especies, respectivamente.
En ambas localidades, las especies más abundantes fueron: Brachidontes adamsianus (Dunker, 1857), Leiosolenus aristatus (Dillwyn, 1817) e Isognomon janus Carpenter, 1857, con 376, 205 y 124 individuos, respectivamente, lo cual representó 91% del total de individuos. Brachidontes adamsianus obtuvo la mayor abundancia en El Palmar a lo largo del ciclo anual, con el valor más elevado (161 individuos) en enero. En Las Gatas, Leiosolenus aristatus fue la especie con la mayor abundanciacon 140 individuos en enero; sin embargo, la especie más abundante fue B. adamsianus a lo largo de los otros meses de muestreo. Mientras que las especies menos abundantes fueron: Sphenia fragilis (H. Adams & A. Adams, 1854) con 2 individuos en enero y mayo en El Palmar y 1 individuo en noviembre en Las Gatas, Crassinella ecuadoriana Olsson, 1961 con 2 individuos en mayo y julio en El Palmar, Linucula declivis (Hinds, 1843) con 2 individuos en julio y noviembre en Las Gatas, Mya sp. y Mytilus edulis Linnaeus, 1758 con 1 individuo en enero en El Palmar (tabla 1).
Tabla 1
Abundancia de las especies de bivalvos recolectadas en las localidades de estudio por mes de muestreo. El arreglo sistemático sigue la propuesta de Horton et al. (2024).
Especie | Playa El Palmar | Playa Las Gatas | |||||||
ene | may | jul | nov | ene | may | jul | nov | Total | |
Subclase Protobranchia | |||||||||
Orden Nuculida | |||||||||
Familia Nuculidae | |||||||||
Linucula declivis | – | – | – | – | – | – | 1 | 1 | 2 |
Subclase Autobranchia | |||||||||
Orden Mytilida | |||||||||
Familia Mytilidae | |||||||||
Brachidontes adamsianus | 161 | 43 | 18 | 14 | 41 | 9 | 34 | 56 | 376 |
Brachidontes semilaevis | 4 | – | – | 1 | – | – | – | – | 5 |
Leiosolenus aristatus | 2 | 30 | 1 | 6 | 140 | 7 | 5 | 14 | 205 |
Modiolus capax | 1 | 2 | 1 | 1 | – | – | – | – | 5 |
Mytilus edulis | 1 | – | – | – | – | – | – | – | 1 |
Orden Arcida | |||||||||
Familia Arcidae | |||||||||
Acar rostae | – | – | – | – | – | – | 1 | 3 | 4 |
Orden Ostreida | |||||||||
Familia Pteriidae | |||||||||
Isognomon janus | 51 | 32 | 8 | 5 | 2 | – | 10 | 16 | 124 |
Familia Pinnidae | |||||||||
Pinna rugosa | – | – | – | – | – | – | 3 | – | 3 |
Orden Lucinida | |||||||||
Familia Lucinidae | |||||||||
Parvilucina approximata | 1 | – | – | – | – | – | – | – | 1 |
Orden Carditida | |||||||||
Familia Carditidae | |||||||||
Carditamera affinis | 5 | 2 | – | – | 1 | – | – | 1 | 9 |
Carditamera radiata | – | 3 | 1 | – | 3 | – | 3 | 5 | 15 |
Cardites grayi | – | 1 | 3 | – | – | – | – | – | 4 |
Familia Crassatellidae | |||||||||
Crassinella ecuadoriana | – | 1 | 1 | – | – | – | – | – | 2 |
Orden Venerida | |||||||||
Familia Chamidae | |||||||||
Chama coralloides | 3 | 2 | 1 | – | – | – | – | 7 | 13 |
Orden Myida | |||||||||
Familia Myidae | |||||||||
Mya sp. | 1 | – | – | – | – | – | – | – | 1 |
Sphenia fragilis | 1 | 1 | – | – | – | – | – | 1 | 3 |
Total | 231 | 117 | 34 | 27 | 187 | 16 | 57 | 104 | 773 |
Figura 2. Especies de bivalvos asociadas a macroalgas recolectadas en Ixtapa-Zihuatanejo, Guerrero, en vista dorsal y ventral. a, Linucula declivis;b, Brachidontes adamsianus;c, Leiosolenus aristatus; d, Mytilus edulis; e, Acar rostae; f, Isognomon janus;g, Pinna rugosa; h, Parvilucina approximata; i, Carditamera radiata; j, Cardites grayi; k, Crassinella ecuadoriana; l, Chama coralloides;m, Sphenia fragilis; n, Modiolus capax;o, Brachidontes semilaevis;p, Carditamera affinis;q, Mya sp.
La riqueza de especies con relación al ciclo anual fue de 13 especies en playa El Palmar, variando entre 5 (noviembre) y 10 especies (enero) a lo largo de los meses, mientras que en playa Las Gatas se obtuvieron 9 especies, con una variación de 2 (mayo) y 9 (noviembre) especies (tabla 1). La curva de acumulación de especies mostró un comportamiento asintótico en El Palmar, mientras que Las Gatas no mostró una tendencia a ser asintótica (fig. 3).
Figura 3. Curva de acumulación de especies de bivalvos registrados en las localidades de estudio. Línea azul: playa El Palmar, línea verde: playa Las Gatas.
En playa El Palmar el nivel intermareal bajo obtuvo el mayor número de especies (13) y el menor número de especies se presentó tanto en el nivel intermareal medio como en el alto (7); mientras que en playa Las Gatas, se presentó el mayor número de especies en el nivel intermareal medio (10) y el nivel alto tuvo el menor número de especies (5). Las especies que se presentaron en ambas localidades en los 3 niveles de la zona intermareal fueron Brachidontes adamsianus, Leiosolenus aristatus e Isognomon janus.
En El Palmar se registraron valores del índice de diversidad de Shannon que fluctuaron entre 1.30 y 2.15 bits/individuo, mientras que en Las Gatas este índice varió entre 0.98 y 2.09 bits/individuo (tabla 2). En general, el índice de equidad de Pielou presentó valores mayores a 0.60 en playa El Palmar, excepto en enero, mientras que en playa Las Gatas se observó que los valores fueron bajos en enero y mayo, y mayores a 0.60 en julio y noviembre. Las pruebas de “t de student”para el índice de diversidad de Shannon entre los meses de muestreo, indicaron diferencias significativas entre todos ellos (p ≤ 0.05) en El Palmar, y Las Gatas (tabla 3).
Tabla 2
Índices comunitarios calculados para las especies recolectadas en playa El Palmar y playa Las Gatas. N = número de individuos, H’= Índice de Shannon, J’= equidad de Pielou y λ = dominancia de Simpson. Los valores máximos y mínimos por localidad para el índice de Shannon (H´) se marcan con negritas.
Playa El Palmar | Playa Las Gatas | |||||||
Mes | N | H’ | J’ | λ | N | H’ | J’ | λ |
Enero | 230 | 1.30 | 0.48 | 0.54 | 187 | 0.99 | 0.41 | 0.60 |
Mayo | 117 | 2.15 | 0.81 | 0.27 | 16 | 0.98 | 0.41 | 0.50 |
Julio | 34 | 2.03 | 0.77 | 0.34 | 54 | 1.63 | 0.68 | 0.44 |
noviembre | 27 | 1.77 | 0.67 | 0.35 | 104 | 2.09 | 0.87 | 0.33 |
En el análisis de escalamiento multidimensional no métrico (nMDS), a partir de los valores de abundancia de las especies de bivalvos, se observó que las muestras se separan en 3 grupos: un primer grupo dominado por estaciones pertenecientes a playa El Palmar (P), un segundo grupo principalmente con estaciones de playa Las Gatas (G) y uno tercero con estaciones de ambas localidades que probablemente se separa por la dominancia de la especie Brachidontes adamsianus (fig. 4). El vector indica la dirección a través del plano de ordenación en la cual aumentan los valores de abundancia de las especies dominantes, la longitud de la línea indica la cantidad de variación total de cada especie, entonces, si toda la variación se explicara, la línea azul alcanzaría el círculo azul.
Tabla 3
Resultado de la prueba de “t de student”para evaluar diferencias significativas del índice de Shannon entre los meses de muestreo por localidad.
Playa El Palmar | Playa Las Gatas | |||||||
Meses | H´ | t | p | gl | H´ | t | p | gl |
enero vs. mayo | 1.30/2.15 | -0.009 | 324 | 1 | 0.99/0.98 | 0.009 | 169 | 1 |
mayo vs. julio | 2.15/2.03 | 0.0004 | 142 | 0.98/1.63 | -19.84 | 4 | ||
julio vs. noviembre | 2.03/1.17 | 0.023 | 60 | 1.63/2.09 | -3.18 | 94 |
El análisis de Permanova reveló que la composición de bivalvos difiere significativamente entre las diferentes localidades muestreadas (pseudo-F = 5.2373, p = 0.001). No hubo diferencias estadísticamente significativas para los meses de muestreo (pseudo-F=1.431, p = 0.147) y el nivel de marea (pseudo-F = 1.19, p = 0.309) (tabla 4).
Tabla 4
Resultado de los análisis de Permanova para los factores de localidad, mes de muestreo y nivel de marea. GL: Grados de libertad, ms: media suma de cuadrados, perm: permutaciones, * = valores significativos.
Factor | GL | ms | Pseudo-F | p (perm) |
Localidad | 1 | 7,844.9 | 5.2373 | 0.001* |
Res | 57 | 1,497.9 | ||
Total | 58 | |||
Mes | 3 | 2,249.9 | 1.431 | 0.147 |
Res | 55 | 1,572.3 | ||
Total | 58 | |||
Nivel de marea | 2 | 1,915 | 1.1996 | 0.309 |
Res | 56 | 1,596.3 | ||
Total | 58 |
Riqueza y composición de macroalgas. Se encontró un total de 3 phyla, 3 clases, 5 subclases, 14 órdenes, 20 familias, 33 géneros y 59 especies de macroalgas; de las cuales, a cada localidad le pertenecen 37 especies. De las 59 especies totales, 11 fueron Chlorophyta (18%), 42 Rhodophyta (70%) y 7 Heterokontophyta-Phaeophyceae (11%). Entre ambas localidades se compartieron 16 especies. Las familias con mayor número de especies fueron Rhodomelaceae con 12 especies, Corallinaceae con 10 y Ceramiaceae y Dictyotaceae con 5 especies cada una (tabla 5).
Asociación entre grupos morfofuncionales y moluscos. Las macroalgas se clasificaron en 6 grupos morfofuncionales: algas filamentosas, foliosas, foliosas corticadas, filamentosas corticadas, coriáceas y calcáreas articuladas (tabla 5). Los grupos dominantes en las 2 localidades fueron las algas filamentosas, filamentosas corticadas y las calcáreas articuladas. En playa El Palmar se presentó una mayor riqueza de especies (14) de bivalvos al interior de las comunidades algales que contenían el grupo de algas calcáreas articuladas. En playa Las Gatas, los grupos morfofuncionales que presentaron mayor cantidad de especies fueron las algas filamentosas, filamentosas corticadas y las calcáreas articuladas con 10 especies cada uno (tabla 6).
Cobertura de macroalgas, abundancia de moluscos y sedimentos retenidos. La regresión lineal por localidad indicó que para playa El Palmar hay una correlación baja con 24% entre los componentes: cobertura de macroalgas, abundancia de bivalvos y peso de sedimento húmedo; sin embargo, se observó una relación moderada entre los sedimentos retenidos por las macroalgas y la abundancia de bivalvos (R = 0.55 y R2 = 0.24). Para playa Las Gatas la regresión lineal múltiple indicó una correlación baja con 10% entre los factores y una relación moderada entre la cobertura de macroalgas y la abundancia de bivalvos (R = 0.42 y R2 = 0.10). La cobertura de macroalgas, abundancia de bivalvos y los sedimentos retenidos varió según la localidad (fig. 5A-C), mes de muestreo (fig. 5D-F) y nivel de marea (fig. 5G-I). En El Palmar la regresión lineal mostró una relación directamente proporcional entre la abundancia de bivalvos y la cobertura de macroalgas. Los valores más altos de cobertura de macroalgas (0.89 m2) se relacionaron con los valores más altos de abundancia de bivalvos.
Con respecto al mes de muestreo, las regresiones lineales múltiples indicaron una correlación alta con 98% (R = 0.99 y R2 = 0.98), entre la cantidad de sedimento retenido y la abundancia de bivalvos. La regresión lineal múltiple sugirió que la abundancia de bivalvos disminuyó a lo largo del ciclo anual, donde enero presentó la mayor abundancia (de 187-230 individuos) y noviembre la menor (27 individuos). La relación de la abundancia de bivalvos con el peso húmedo de sedimento tuvo un comportamiento parecido, disminuyendo conforme al ciclo anual: en las muestras de enero se reportó mayor sedimento retenido (0.56 kg), mientras que en julio se encontraron menos sedimentos retenidos (0.17 kg).
Figura 4. Análisis de escalamiento multidimensional no métrico (nMDS) de las especies de bivalvos recolectadas en las localidades de estudio. Las líneas azules (vectores) indican las especies abundantes y frecuentes (dominantes). El círculo azul representa la variación de las abundancias.
Tabla 5
Especies de macroalgas asociadas a bivalvos, para cada localidad y grupo morfofuncional. Se marca con * la especie de macroalga presente en la localidad. Localidad: playa El Palmar (P), playa Las Gatas (G). Grupos morfofuncionales de macroalgas: filamentosas (Fil), foliosas (Fol), foliosas corticadas (Foc), filamentosas corticadas (Fic), coriáceas (Cor), calcáreas articuladas (Cal). Especies de bivalvos: 1) Linucula declivis, 2) Brachidontes adamsianus, 3) Brachidontes semilaevis, 4) Leiosolenus aristatus, 5) Modiolus capax, 6) Mytilus edulis, 7) Acar rostae, 8) Isognomon janus, 9) Carditamera affinis, 10) Carditamera radiata, 11) Cardites grayi, 12) Crassinella ecuadoriana, 13) Chama coralloides, 14) Mya sp., 15) Sphenia fragilis,16) Pinna rugosa, 17) Parvilucina approximata.
Especies | Grupo morfofuncional | Especies de bivalvos | |
P | G | ||
Chlorophyta | |||
Ulvales | |||
Ulvaceae | |||
Ulva californica Wille, 1899 | Fol | * | – |
Ulva intestinalis Linnaeus, 1753 | Fol | – | 2, 4, 13 |
Ulva linza Linnaeus, 1753 | Fol | – | 2, 4, 13 |
Bryopsidales | |||
Bryopsidaceae | |||
Bryopsis pennata var. minor J. Agardh, 1887 | Fil | – | 2, 4, 8 |
Caulerpaceae | |||
Caulerpa chemnitzia (Esper) J.V. Lamouroux, 1809 | Fil | 2, 4, 8, 10, 17 | 1, 2, 4, 7, 8, 9, 10, 15 |
Caulerpa sertularioides (S.G. Gmelin) M. Howe, 1905 | Fil | 2, 4, 8, 10, 17 | 1, 2, 4, 7-10, 15 |
Halimedaceae | |||
Halimeda discoidea Decaisne, 1842 | Cal | 2, 10, 17 | – |
Cladophorales | |||
Cladophoraceae | |||
Chaetomorpha antennina (Bory) Kützing, 1847 | Fil | 2, 3, 7, 8-13 | 4, 7, 8, 11, 13 |
Cladophora sp. | Fil | 2, 4, 5, 8, 9 | – |
Cladophora graminea Collins, 1909 | Fil | – | 16 |
Cladophora microcladioides Collins, 1909 | Fil | 2, 4, 5, 8, 9 | – |
Rhodophyta | |||
Gigartinales | |||
Cystocloniaceae | |||
Hypnea pannosa J. Agardh, 1847 | Fic | 2-6, 8-15 | 2, 4, 8-10, 16 |
Hypnea spinella (C. Agardh) Kützing, 1847 | Fic | 2-6, 8-15, 17 | 2, 4, 8-10, 16 |
Hypnea johnstonii Setchell y N. L. Gardner, 1924 | Fic | – | 1, 2, 4, 7-10, 16 |
Phyllophoraceae | |||
Ahnfeltiopsis gigartinoides (J. Agardh) P. C. Silva y DeCew, 1992 | Fic | 8 | – |
Gymnogongrus johnstonii (Setchell y N. L. Gardner) E.Y. Dawson, 1961 | Fic | 2, 8 | – |
Ceramiales | |||
Ceramiaceae | |||
Centroceras clavulatum (C. Agardh) Montagne, 1846 | Fil | 2-5, 8 | 2, 4, 10, 15 |
Tabla 5. Continúa | |||
Especies | Grupo morfofuncional | Especies de bivalvos | |
P | G | ||
Ceramium sp. | Fil | – | 1, 2, 4, 7, 8, 10, 13, 16 |
Ceramium camouii E. Y. Dawson, 1944 | Fil | – | 1, 2, 4, 7, 8, 10, 13, 16 |
Ceramium zacae Setchell y N. L. Gardner, 1937 | Fil | – | 1, 2, 4, 7, 8, 10, 13, 16 |
Gayliella flaccida (HarveyyKützing) T.O. McIvory y L.J. Cho, 2008 | Fic | – | 2, 4, 7, 8, 11 |
Delesseriaceae | |||
Taenioma perpusillum (J. Agardh) J. Agardh, 1863 | Fil | – | 4, 16 |
Rhodomelaceae | |||
Chondria sp. | Fic | * | * |
Herposiphonia secunda (C. Agardh) Ambronn, 1880 | Fic | 8 | 2, 4, 8, 10 |
Herposiphonia tenella (C. Agardh) Ambronn, 1880 | Fic | 7 | – |
Melanothamnus simplex (Hollenberg) Díaz-Tapia y Maggs, 2017 | Fic | 2, 8 | – |
Melanothamnus sphaerocarpus (Borgesen) Díaz.Tapioa y Maggs, 2017 | Fic | – | 2, 4, 8 |
Polysiphonia mollis J. D. Hooker y Harvey, 1847 | Fic | * | 2, 4, 8, 15 |
Polysiphonia nathanielii Hollenberg, 1958 | Fic | – | 2, 4, 8, 15 |
Polysiphonia subtilissima Montagne, 1840 | Fic | – | 2, 4, 8, 15 |
Eutrichosiphonia confusa (Hollenberg) Savoie y G.W. Saunders, 2019 | Fic | – | 2, 4, 8 |
Laurencia sp. | Fic | 2, 8, | 2, 7, 8, 13 |
Laurencia hancockii E.Y. Dawson, 1944 | Fic | 2, 8 | – |
Laurencia subcorymbosa E.Y. Dawson, 1963 | Fic | 2, 8 | – |
Rhodymeniales | |||
Rhodymeniaceae | |||
Tayloriella dictyurus (J. Agardh) Kylin, 1956 | Fil | 2, 8 | – |
Lomentariaceae | |||
Ceratodictyon tenue (Setchell y N. L. Gardner) J.N. Norris, 2014 | Fil | – | 2, 4, 8 |
Corallinales | |||
Corallinaceae | |||
Amphiroa beauvoisii J.V. Lamouroux, 1816 | Cal | 2-15, 17 | 1, 2, 4, 7-14, 16 |
Amphiroa misakiensis Yendo, 1902 | Cal | 2-15 | 1, 2, 4, 7-14, 16 |
Amphiroa rigida J.V. Lamouroux, 1816 | Cal | 2-15 | 1, 2, 4, 7-14, 16 |
Amphiroa subcylindrica E.Y. Dawson, 1953 | Cal | 2-15 | – |
Jania capillacea Harvey, 1853 | Cal | – | 2, 4, 7-10, 15, 16 |
Jania subpinnata E.Y. Dawson, 1953 | Cal | 2, 4, 5, 8, 9, 12, 14 | 2, 4, 7-10, 15, 16 |
Jania tenella (Kützing) Grunow, 1874 | Cal | 2, 4, 5, 8, 9, 12, 14 | 2, 4, 7-10, 15, 16 |
Jania tenella var. tenella | Cal | 2, 4, 5, 8, 9, 12, 14 | – |
Gelidiales | |||
Gelidiaceae | |||
Gelidiella acerosa (Forsskal) Feldmann y Hamel, 1934 | Fic | – | 2, 7, 8 |
Gelidium mcnabbianum (E.Y. Dawson) B. Santelices, 1998 | Fic | – | 2, 4, 7-11 |
Gelidium pusillum (Stackhouse) Le Jolis, 1863 | Fic | 2, 4, 5, 8, 12, 14 | 2, 4, 7-11 |
Pterocladiaceae | |||
Pterocladiella caloglossoides (M. Howe) Santelices, 1998 | Fic | – | 2, 4, 7, 8, 15 |
Gracilariales | |||
Gracilariaceae | |||
Gracilaria sp. | Fic | 2, 8, 12, 14 | – |
Halymeniales | |||
Halymeniaceae | |||
Grateloupia huertana Mateo-Cid, Mendoza-González y Gavio, 2005 | Fic | – | 4, 8 |
Grateloupia versicolor (J. Agardh) J. Agardh, 1847 | Fic | 2, 4, 8 | – |
Heterokontophyta-Phaeophyceae | |||
Ectocarpales | |||
Scytosiphonaceae | |||
Chnoospora minima Papenfuss, 1956 | Fic | * | – |
Fucales | |||
Sargassaceae | |||
Sargassum liebmannii Agardh, 1847 | Cor | 2, 4, 5, 7, 8, 10-13, 15, 17 | – |
Dictyotales | |||
Dictyotaceae | |||
Dictyota sp. | Foc | – | * |
Dictyota dichotoma (Hudson) J.V. Lamouroux, 1809 | Foc | 2, 4, 5, 8 | – |
Lobophora variegata (J.V.Lamouroux)Womersley ex E.C.Oliveira, 1977 | Fic | 8 | – |
Padina mexicana var. erecta Avila-Ortiz, 2003 | Foc | 2, 4, 5, 8, 9 | – |
Padina ramonribae Avila-Ortiz, Pedroche y Díaz-Martínez, 2016 | Foc | 2, 4, 5, 8, 9 | – |
La regresión lineal por nivel de marea indicó una correlación significativa con 100% (R = 1 y R2 = 1), es decir, la cobertura de algas y el sedimento retenido están relacionados con la abundancia de bivalvos en los diferentes niveles de marea. El nivel del intermareal bajo presentó los valores más elevados de cobertura de macroalgas con 0.61 m2 (fig. 5G), mayor abundancia de moluscos con 275 individuos (fig. 5H) y mayor cantidad de sedimentos con 0.69 kg (fig. 5I). En el intermareal alto se registraron los valores más bajos de abundancia de bivalvos con 230 individuos (fig. 5H), cobertura de algas con 0.52 m2 (fig. 5G) y peso húmedo de sedimentos con 0.30 kg (fig. 5I), es decir, conforme se asciende en el intermareal, estas variables disminuyen.
Tabla 6
Asociación de las especies de bivalvos con los grupos morfofuncionales propuestos por Steneck y Dethier (1994). Grupos morfofuncionales de macroalgas: filamentosas (Fil), foliosas (Fol), foliosas corticadas (Foc), filamentosas corticadas (Fic), coriáceas (Cor), calcáreas articuladas (Cal). × = Presencia de la especie.
Familia | Especie | Playa El Palmar | Playa Las Gatas | ||||||||||
Fil | Fol | Foc | Fic | Cor | Cal | Fil | Fol | Foc | Fic | Cor | Cal | ||
Nuculidae | L. declivis | – | – | – | – | – | – | × | – | – | × | – | × |
Mytilidae | B. adamsianus | × | – | × | × | × | × | × | × | – | × | – | × |
B. semilaevis | × | – | – | × | – | × | – | – | – | – | – | – | |
L. aristatus | × | – | × | × | × | × | × | × | – | × | – | × | |
M. capax | × | – | × | × | × | × | – | – | – | – | – | – | |
M. edulis | – | – | – | × | – | × | – | – | – | – | – | – | |
Arcidae | A. rostae | × | – | – | – | × | × | × | – | – | × | – | × |
Pteriidae | I. janus | × | – | × | × | × | × | × | – | – | × | – | × |
Pinnidae | P. rugosa | – | – | – | – | – | – | × | – | – | × | – | × |
Lucinidae | P. approximata | × | – | – | × | × | × | – | – | – | – | – | – |
Carditidae | C. affinis | × | – | × | × | – | × | × | – | – | × | – | × |
C. radiata | × | – | – | × | × | × | × | – | – | × | – | × | |
C. grayi | × | – | – | × | × | × | × | – | – | × | – | × | |
Crassatellidae | C. ecuadoriana | × | – | – | × | × | × | – | – | – | – | – | – |
Chamidae | C. coralloides | × | – | – | × | × | × | × | × | – | × | – | × |
Myidae | Mya sp. | – | – | – | × | – | × | – | – | – | – | – | – |
S. fragilis | – | – | – | × | × | × | × | – | – | × | – | × | |
Total | 12 | 0 | 5 | 14 | 11 | 15 | 11 | 3 | 0 | 11 | 0 | 11 |
Discusión
Riqueza y composición de moluscos. De las 17 especies encontradas en el presente estudio, 16 ya habían sido observadas para el Pacífico tropical mexicano, específicamente para Ixtapa-Zihuatanejo en sustratos rocosos y arenosos (Flores-Rodríguez et al., 2007, 2012; Lesser, 1984; López-Rojas et al.,2017; Lozada, 2010; Sánchez, 2014), excepto Mytilus edulis que es un nuevo registro para Guerrero. Esta especie fue encontrada en playa El Palmar, asociada a los géneros de macroalgas Amphiroa e Hypnea. Dicha especie había sido registrada para el océano Pacífico desde las costas del Ártico, Canadá, EUA, hasta Cabo San Lucas, México, al interior del golfo de California en San Luquitas y Santa Rosalía, y para el Pacífico tropical mexicano en isla Socorro, por lo que su distribución se amplía hacia el sur, a Guerrero en ambientes litorales asociados a las comunidades de macroalgas (Cadena-Cárdenas et al., 2009; Fitch, 1953; Shaw et al., 1988). La ampliación de la distribución de esta especie puede deberse al movimiento de la corriente de California como ha sido reportado para bivalvos (Schulien et al., 2020), gasterópodos y escafópodos que ampliaron su distribución desde la provincia Californiana hacia la Panámica donde se señala la confluencia de especies de moluscos entre ambas provincias (Landa-Jaime y Arciniega-Flores 1998; Ríos-Jara et al., 2003; Gama-Kwick et al., 2021). Otra de las razones por la cual M. edulis amplió su distribución al sur podría ser por el fenómeno de El Niño que proporciona condiciones ambientales para que las especies que habitan en sitios templados y subtropicales se desplacen a sitios tropicales (Díaz y Ortlieb, 1993; Paredes et al., 1998; Velez y Zeballos, 1985).
Figura 5. Variación de cobertura de macroalgas, abundancia de bivalvos y sedimento retenido por localidades (A, B y C), mes de muestreo (D, E y F) y nivel de la zona intermareal (G, H e I). Verde = macroalgas, amarillo = bivalvos, rojo = sedimento retenido.
La familia mejor representada fue Mytilidae con 5 especies: Brachidontes adamsianus, B. semilaevis, Leiosolenus aristatus, Modiolus capax y Mytilus edulis (tabla 1), dichas especies coinciden con lo registrado por Flores-Garza et al.(2014), Galeana-Rebolledo et al.(2012) y Garcés (2011) en Acapulco, así como por López-Rojas et al. (2017) en diferentes localidades en Guerrero. Mytilidae también ha sido la familia mejor representada en otras localidades de Zihuatanejo (muelle municipal), donde se han observado 8 especies (Guzmán, 2022), así como en otras regiones tropicales como Brasil, donde esta familia también fue la mejor representada con 5 especies (Santos et al., 2020).
Una mayor riqueza de especies de la familia Mytilidae en Ixtapa-Zihuatanejo podría explicarse debido a la frecuencia y rápido asentamiento de larvas de bivalvos, ya que son organismos que están en constante reproducción a lo largo del año (Seed, 1969a). Suchanek (1978) mencionó que los miembros de la familia Mytilidae tienden a colonizar rápidamente los espacios disponibles en la zona intermareal rocosa, familia que se ha adaptado a distintos hábitats (Keen, 1971), gracias a su rápido aumento de talla y posterior asentamiento de larvas (Ceccherelli y Rossi, 1984). También se ha mencionado que los organismos de dicha familia pueden producir fibras de biso, por lo que son capaces de anclarse y permanecer en sustratos en los que otras familias de bivalvos no pueden, como rocas, arenas u otros organismos (Keen, 1971; Stella et al.,2010).
En general, se encontró una mayor riqueza de especies en El Palmar, valores que pueden explicarse debido a la heterogeneidad ambiental de una playa expuesta con oleaje intenso como dicho sitio (Morales et al., 2008). Se ha mencionado que la riqueza de especies de bivalvos es mayor en zonas expuestas en donde las olas impactan directamente, por su parte, las zonas protegidas o de menor oleaje presentan menor riqueza específica (Flores-Garza et al., 2014; Flores-Rodríguez et al., 2012; Valdés-González et al., 2004).
En El Palmar, se encontraron valores elevados para la riqueza de especies de bivalvos (17 spp.) respecto de lo encontrado en Las Gatas, lo anterior también se ha observado para otras clases de moluscos como los gasterópodos y poliplacóforos asociados a comunidades de macroalgas en playa El Palmar (Aguilar, 2017). También en esta playa se encontró menor cantidad de sedimentos retenidos por las macroalgas (fig. 5C). El Palmar está conformada por rocas de diferentes tamaños, que pueden moverse conforme se dan los cambios en el oleaje y la marea, lo que le confiere un alto dinamismo que no permite la acumulación de sedimentos (Gibbons, 1988), tal como se observó en el presente estudio. Dichas características ayudan a que se conformen una amplia gama de ambientes que pueden ser colonizados, donde se esperaría una mayor riqueza de especies como se observó para los bivalvos de El Palmar (Benedetti-Cecchi, 2001).
Como se ha mencionado anteriormente, playa Las Gatas es un sitio protegido, por esta razón es un lugar de baja energía, con características poco favorables para el establecimiento de especies de bivalvos de manera muy similar a los estudios realizados en otras zonas de Guerrero (Flores-Garza et al., 2014; Flores-Rodríguez et al., 2012; Valdés-González et al., 2004). Las fluctuaciones observadas entre ambas playas de la riqueza de especies, también podrían deberse a otro tipo de características presentes en algunas familias de bivalvos que no tienen hábitos epifaunales como los que se adhieren a las macroalgas y que presentan hábitos de vida semiinfaunales o infaunales (Garcés, 2011), por lo que estas especies podrían encontrarse en otros sitios, posiblemente en zonas más profundas en la zona submareal, o de igual forma, las especies de bivalvos se pueden ver afectadas por la contaminación al incorporar en sus tejidos bacterias patógenas (Gosling, 2015), que es común de una zona turística como Ixtapa-Zihuatanejo (IMTA, 2010; UNAM, 2013). Lo anterior podría explicar las fluctuaciones en la riqueza de especies de bivalvos observadas en Zihuatanejo.
Los valores de riqueza de especies no variaron considerablemente en las localidades estudiadas a lo largo del año. La riqueza de especies de bivalvos en un ciclo anual puede ser constante, ya que la mayoría de las especies (70%) de bivalvos del presente trabajo son generalmente inmóviles o sedentarias. Lo anterior hace que sus poblaciones se mantengan con pocos cambios con respecto a su riqueza de especies debido a sus patrones de reproducción anuales o bianuales (Baqueiro y Masso, 1988; Flores-Rodríguez et al., 2012; Seed, 1969a).
La especie más abundante en el presente trabajo fue B. adamsianus (Mytilidae). Esta especie ha sido registrada como la más abundante en otras localidades de Zihuatanejo (Guzmán, 2022). En particular, las familias Chamidae y Mytilidae se han registrado como las más abundantes para Guerrero; por su parte, la especie Isognomon janus Carpenter, 1857 (Isognomonidae) también ha presentado un gran número de organismos (Guzmán, 2022; López-Rojas et al. 2017). Las bajas abundancias de los bivalvos encontrados en este trabajo pueden estar dadas por el tipo de hábitos de vida de dichos moluscos. Por ejemplo, el género Sphenia tienen un hábito endolítico (Esqueda-González et al., 2014; Garcés, 2011; Guzmán, 2022), por lo que sería poco probable encontrarlos dentro de las comunidades de macroalgas, ya que es común encontrarlos incrustados en agujeros preexistentes de rocas, fragmentos de madera, incluso otros materiales como las conchas de moluscos, así como entre colonias de briozoos (Coan, 1999). El Permanova mostró que existen diferencias significativas entre localidades, esto puede deberse a que playa Las Gatas es considerada como un sitio protegido y playa El Palmar como un sitio expuesto, esto representó diferencias en abundancia y riqueza de especies, que pueden deberse a los requerimientos medioambientales propios de cada una de las especies de bivalvos como: sustrato, alimento, salinidad o temperatura (Borges et al., 2014; Galeana-Rebolledo et al., 2012; Seed 1969a).
En playa El Palmar, la curva de acumulación de especies fue asintótica, lo que sugiere que se encontró a la mayoría de los bivalvos asociados a macroalgas de esta localidad. Mientras que, en playa Las Gatas, la curva de acumulación de especies no fue asintótica. La cantidad de especies recolectadas en un sitio está relacionada con el esfuerzo de muestreo (Moreno, 2001). Por lo que el menor número de muestras en donde se encontraron bivalvos en playa Las Gatas (25), en comparación con las de playa El Palmar (35), podría explicarse por este hecho. Se ha mencionado que una curva de acumulación de especies en muy raras ocasiones llega a ser asintótica, ya que siempre habrá especies que no se recolecten o durante los muestreos pueden encontrarse especies raras, lo que puede estar determinado por el sitio de muestreo, temporada del año, tipo de sustrato, entre otras variables (Jiménez-Valverde y Hortal, 2003).
Además, otra de las razones para que la curva de acumulación de especies no haya sido asintótica para Las Gatas, podría deberse a que, en ambas localidades, las muestras de bivalvos fueron obtenidas de comunidades de macroalgas, que es un sustrato muy específico y es distinto a las rocas y arena que componen la zona intermareal que corresponden a los sustratos que se han estudiado en la mayoría de los distintos trabajos malacológicos de Guerrero. Por lo anterior, el inventario podría estar incompleto; sin embargo, es una buena aproximación de la biodiversidad de las especies de bivalvos asociadas a comunidades de macroalgas de la zona norte de Guerrero en Ixtapa-Zihuatanejo. Para tener una curva de acumulación de especies asintótica se recomienda muestrear todos los sustratos posibles y de esta forma, obtener casi todas las especies de bivalvos de la zona.
Se encontraron cambios con respecto a la abundancia de bivalvos asociados a macroalgas a lo largo del año, fluctuaciones que pudieran estar relacionadas con sus patrones reproductivos, lo que podría explicar la disminución en la abundancia de bivalvos en los meses cálidos (mayo y julio) y un aumento en la cantidad de individuos en los meses fríos (enero y noviembre). Los cambios en la abundancia de bivalvos se han estudiado en trabajos sobre reproducción de diferentes especies de moluscos (Baqueiro y Aldana, 2000, 2003). Algunas de las especies de bivalvos que han sido utilizadas en estudios reproductivos o sobre ciclos gonádicos son: Chione undatella (G. B. Sowerby I, 1835), Megapitaria aurantiaca (G. B. Sowerby I, 1831) (Veneridae), Brachidontes rodriguezii (d’Orbigny, 1842), Mytilus edulis y Mytilus chilensis Hupé, 1854 (Mytilidae), así como Anadara tuberculosa (G. B. Sowerby I, 1833) (Arcidae), tanto en México como otros países de Sudamérica. Dichos estudios destacan que los bivalvos se reproducen continuamente a lo largo del año y se ha mencionado que el momento del desove comprende desde la primavera hasta el otoño con temperaturas por encima de los 25 °C; se ha observado que las larvas se instalan en octubre y noviembre, con una metamorfosis de 15 a 25 días, dichas larvas permanecen adheridas a las algas de noviembre a mayo, en donde llevan a cabo su metamorfosis (Aguillón, 2011; Baqueiro y Masso, 1988; García-Domínguez et al., 2008; Hernández-Moreno et al., 2020; Oyarzún et al., 2011; Seed, 1969a; Torroglosa, 2015). Solo se encontró a Mytilus edulis como especie de importancia comercial para el presente trabajo. En especies que no se les ha considerado importantes para la industria pesquera, es muy poco lo que se conoce acerca de sus hábitos reproductivos (Aguilar, 2017).
Las 3 especies más abundantes del presente trabajo son epifaunales de hábitos filtradores por suspensión (García-Cubas, 1981), Isognomon janus (Isognomonidae) habita en la zona intermareal hasta profundidades de 20 m, Leiosolenus aristatus (Mytilidae)puede encontrarse desde el litoral hasta los 300 m sobre rocas u otros bivalvos (Coan y Valentich-Scott, 2012) y Brachidontes adamsianus (Mytilidae)habita sobre grietas de rocas grandes en zonas expuestas (Landa-Jaime et al., 2013). Las especies que no pertenecen a las familias Isognomonidae y Mytilidae, en general, son organismos infaunales de hábitos filtradores por suspensión, a excepción de la familia Myidae que son infaunales perforadores de hábitos filtradores por suspensión (García-Cubas, 1981; Coan, 1999).
Los bivalvos están presentes en diferentes lugares al interior de la zona intermareal rocosa y hay especies que tienen preferencia por algún sitio a lo largo de dicha zona (Román-Contreras et al.,1991; Sibaja-Cordero y Vargas-Zamora, 2006; Suchanek, 1978). En el presente estudio se observó un incremento en la abundancia y riqueza en los niveles intermareales medio y bajo. Este mismo patrón se ha observado en otras especies de moluscos como gasterópodos y poliplacóforos de Zihuatanejo (Aguilar, 2017; Gama-Kwick et al., 2021).
El índice de diversidad de Shannon obtenido en este trabajo fue bajo, menor a 2.16 bits/individuo para ambos sitios de muestreo, en comparación con otros trabajos sobre bivalvos asociados a macroalgas para Zihuatanejo, como el de Guzmán (2022), donde se encontraron valores de 3.41 bits/individuo en el muelle municipal. Flores-Garza et al.(2014), en su estudio en Acapulco, presentaron valores de 3.65 bits/individuo y Galeana-Rebolledo et al.(2012) reportaron valores de 3.64 bits/individuo. Las diferencias con respecto al índice de diversidad de Shannon pueden deberse a la baja riqueza de especies encontrada en Ixtapa-Zihuatanejo, respecto de otras localidades del Pacífico tropical mexicano, donde se han registrado valores superiores a 17 especies; Reguero y García-Cubas (1989) encontraron 53 especies de bivalvos para Nayarit, y Esqueda-González et al. (2014) reportaron 89 especies de bivalvos para Sinaloa. Lo anterior también puede relacionarse con los altos valores de abundancia de ciertas especies, como es el caso de Brachidontes adamsianus. Ésto también se ve reflejado en los valores calculados del índice de dominancia de Simpson, el cual dio como resultado valores moderados, ya que la abundancia de las especies de bivalvos no es homogénea a lo largo de la zona intermareal rocosa en ambas localidades del presente estudio.
Garcés (2011), describió valores similares del índice de diversidad de Shannon para especies de bivalvos en sustratos rocosos en Acapulco (2.41 bits/individuo) y mencionó que éstos varían según el tipo de sustrato, ya que la riqueza de especies fue mayor en sustratos arenosos que en sustratos rocosos, lo que ocasiona que en estos sustratos, los valores disminuyan.
En el nMDS se observan 3 agrupaciones: una con muestras de la localidad de playa El Palmar, otra con estaciones de playa Las Gatas y la tercera con estaciones de ambas localidades, donde solamente hubo presencia de la especie Brachidontes adamsianus (tabla 1). Dicha especie es una de las más comunes en varias localidades de Jalisco, Oaxaca y Guerrero, por lo que sus valores de abundancia en Ixtapa-Zihuatanejo concuerdan con lo encontrado en otras contribuciones (Castro-Mondragón et al., 2016; Galeana-Rebolledo et al., 2012; Garcés, 2011; Holguín-Quiñones y González-Pedraza 1989; Landa-Jaime et al., 2013; López-Rojas et al., 2017; Torreblanca-Ramírez et al., 2012).
Asociación entre grupos morfofuncionales y moluscos. Los bivalvos recolectados tuvieron una mayor presencia en el grupo morfofuncional de algas filamentosas en ambas localidades, este mismo resultado se observó en muelle municipal en Zihuatanejo por Guzmán (2022) para los bivalvos asociados a macroalgas. Lo anterior puede verse favorecido porque este grupo de algas proporciona refugio ante el oleaje y la desecación, de igual forma tienen un papel fundamental en el asentamiento de las larvas de bivalvos, ya que son utilizadas como un sitio para evitar la competencia entre estadios juveniles y adultos (Dobretsov, 1999; Seed, 1969a).
Las larvas de la familia Mytilidae son atraídas hacia los filamentos de algas rojas de los géneros Ceramium y Polysiphonia, en donde llevan a cabo su metamorfosis (Seed, 1969a), ambos géneros fueron encontrados en el presente trabajo. Las especies de bivalvos de la familia Mytilidae también se encontraron junto con otros géneros de macroalgas como: Bryopsis, Caulerpa, Chaetomorpha y Cladophora (Chlorophyta), dichos géneros también se han relacionado a comunidades de bivalvos en el Caribe colombiano (Quirós-Rodríguez y Campos, 2013). Los géneros Gayliella, Herposiphonia, Taenioma y Tayloriella (Rhodophyta) también se han observado junto a especies de bivalvos, en particular el género Tayloriella se ha encontrado vinculado a 9 diferentes especies de bivalvos en el muelle municipal en Zihuatanejo (Guzmán, 2022). Aun cuando es poco lo que se conoce sobre las asociaciones de bivalvos con macroalgas, es posible sugerir que las larvas de las otras familias de bivalvos encontradas en el presente trabajo (Arcidae, Carditidae, Crassatellidae, Chamidae, Myidae y Pteriidae) puedan tener un comportamiento similar al de Mytilidae y, por esta razón, se podría reconocer su presencia en las comunidades de macroalgas.
La asociación de las especies de bivalvos (B. adamsianus, B. semilaevis, L. aristatus, M. capax, M. edulis, A. rostae, I. janus, C. affinis, C. radiata, C. grayi, C. ecuadoriana, C. coralloides, Mya sp. y S. fragilis) con el grupo de algas filamentosas corticadas fue explicado para especies del género Gigartina porSeed (1969a), quien afirmó que este tipo de algas proporciona una mayor protección y partículas de sedimento retenido (alimento), comparado con las algas filamentosas. En el presente trabajo se encontraron especies de los géneros Ceratodyction, Gelidiella, Gelidium, Gracilaria e Hypnea, que podrían tener una relación similar a la dada por las especies de Gigartina, ya que comparten el mismo grupo morfofuncional propuesto por Steneck y Dethier (1994).
Todas las especies de bivalvos se asociaron al grupo de algas calcáreas articuladas. Las especies del género Corallina proporcionan condiciones ambientales adecuadas para los bivalvos como: refugio, captación de partículas y cantidad de CaCO3 (Seed, 1969a). De esta forma, los bivalvos son capaces de alcanzar tallas mayores en menor tiempo (Seed, 1969a, b). Probablemente, la cantidad de CaCO3 obtenida del ambiente por las macroalgas, queda disponible para que los bivalvos la aprovechen, una vez que las macroalgas mueren. En el presente trabajo, se observó a las especies de algas calcáreas articuladas de los géneros Amphiroa, Halimeda y Jania asociadas a las especies de bivalvos son muy similares a las del género Corallina y comparten el mismo grupo morfofuncional, por lo que podrían funcionar de la misma forma al ser algas que pueden fijar CaCO3 y ser una fuente de aprovechamiento para otros organismos; estas algas calcáreas son la fuente principal de carbonatos marinos (Feely et al., 2004).
Cobertura de macroalgas, abundancia de moluscos y sedimento retenido. La cobertura de macroalgas, abundancia de bivalvos y sedimento retenido disminuyeron conforme se aumentó el nivel del intermareal, los valores más elevados para dichas variables se observaron en el nivel del intermareal bajo (fig. 5G-I). Los resultados de la regresión lineal múltiple sustentan lo anterior, ya que se encontró una relación entre la cobertura de macroalgas y abundancia de bivalvos. Lo anterior podría explicarse por las características del terreno (emersión), debido a que, en un ambiente cambiante como la zona intermareal rocosa, las macroalgas quedan expuestas a una mayor radiación solar en el nivel alto. Por lo tanto, al haber menos céspedes algales (Huovinen et al., 2006), la retención de sedimento es baja, estos 2 factores influyen en las necesidades de cada especie de bivalvo, lo que determina la supervivencia de estos moluscos en esta zona (Airoldi, 2003; Rosenberg, 1977).
Renaud et al. (1997) encontraron una relación entre la abundancia de macroalgas con respecto al sedimento, donde la cobertura de macroalgas era sistemáticamente mayor en las zonas con baja cantidad de sedimentos y un aumento notable en la abundancia de macroalgas se observó después de la remoción de sedimentos. Esto se relaciona con lo encontrado en el presente trabajo, donde se observó una menor cobertura de macroalgas asociada a valores elevados de sedimento. Con respecto de los bivalvos, Forster y Zettler (2004) observaron que la biomasa de Mya arenaria Linnaeus, 1758 se redujo con la presencia de sedimentos finos, lo que podría sugerir que en playa Las Gatas existe una menor abundancia de bivalvos (Airoldi, 2003).
El sedimento en las comunidades de macroalgas queda retenido ya que se acumula en los espacios que hay entre los talos; las cantidades de sedimento acumuladas pueden estar determinadas por la complejidad estructural del alga y las condiciones del medio (García, 2009). La variación en la tasa de sedimentación podría ser un factor de alteración y estrés sobre las comunidades de macroalgas, lo que ocasiona que la tasa de crecimiento algal se vea reducida por la falta de incidencia de luz. Por ello, la fauna que ahí se establece, como las especies de bivalvos, también es menor, ya que una gran cantidad de sedimentos retenidos provoca que los organismos puedan sofocarse y el reclutamiento larval se vea reducido (Airoldi, 2003; Rosenberg, 1977). La clase Bivalvia tiene una preferencia por sustratos arenosos, rocosos, lodosos y fangosos, sin embargo, cuando los sitios no tienen corrientes fuertes y la acumulación de sedimento es alta, la abundancia de bivalvos baja drásticamente (Gosling, 2015).
Las larvas de bivalvos se instalan sobre las algas, donde permanecen adheridas a ellas y llevan a cabo su metamorfosis entre 15 y 25 días; en este lapso, los céspedes algales les proporcionan alimento por el sedimento que retienen y protección contra depredadores y factores abióticos. Posteriormente, los bivalvos juveniles migran fuera de las algas ya que han alcanzado una talla óptima o las algas ya no les brindan suficientes recursos, y llegan a otro sitio que será el definitivo para llevar a cabo el resto de su ciclo de vida (Seed, 1969a, 1969b; Suchanek, 1978).
Olafsson (1986) encontró una relación significativa entre la abundancia de bivalvos y el sedimento, estos resultados son consistentes con lo encontrado en el presente estudio en Ixtapa-Zihuatanejo. Sin embargo, Vázquez (2009) en playa Las Gatas no encontró relación entre la riqueza de especies y el sedimento retenido. Por su parte, en playa El Palmar, que es una zona expuesta, una posible explicación para la relación entre abundancia de bivalvos y sedimento es debido al tipo de hábitos de vida de la mayoría de las especies encontradas en el presente trabajo, ya que la mayoría de ellas son epifaunales o infaunales con hábitos filtradores suspensívoros. La erosión del sustrato rocoso por la acción de las olas y por el movimiento de los sedimentos puede generar heterogeneidad ambiental (Airoldi, 2003). La mezcla constante del agua provee mayor humedad, así como nutrientes al suspender los sedimentos, lo que ocasiona una mayor supervivencia de los organismos, lo que explicaría la relación entre abundancia y sedimento (Gama-Kwick et al., 2021).
Las condiciones de playa El Palmar con un oleaje intenso (Morales et al., 2008), no permite la retención de sedimentos, su baja cantidad podría influir en la composición de las macroalgas. Por lo tanto, la composición de los invertebrados al interior de éstos puede ser muy particular en esta localidad (Gama-Kwick et al., 2021), lo que puede incrementar su diversidad (Chemello y Milazzo, 2002; Prathep et al., 2003). En contraste, playa Las Gatas es un sitio protegido con un oleaje de baja energía y alta cantidad de sedimentos, por lo que su comunidad de macroalgas y bivalvos puede ser menos diversa, abundante y compleja. Lo anterior es similar a lo encontrado en trabajos de la zona para otras clases de moluscos como gasterópodos y poliplacóforos (Aguilar, 2017).
Las comunidades de macroalgas pueden funcionar como sustrato de captura de las larvas desde la columna de agua (Aguilar, 2017). Dichos sitios constituyen zonas de crianza en parte o en todo el ciclo de vida de los bivalvos, ya que proporcionan protección y alimento (Seed, 1969a, b), que está determinado por los sedimentos que las macroalgas retienen.
El presente trabajo aporta resultados que contribuyen al conocimiento de la biodiversidad marina, en especial de los bivalvos asociados a la ficoflora del Pacífico tropical mexicano. Los esfuerzos posteriores deberían centrarse en explorar la biodiversidad de los invertebrados sobre sustratos diferentes en el litoral rocoso, posiblemente utilizando metodologías similares a las del presente estudio con la finalidad de generar comparaciones. En los estudios futuros se debe profundizar en el conocimiento de la distribución geográfica de Mytilus edulis y así corroborar los resultados de este trabajo. Se debe promover el estudio de las interacciones ecológicas, ya que muchas especies de moluscos son impactadas por diferentes factores tanto bióticos, como la depredación o la epibiosis (Aguilar-Estrada et al., 2022; García-Ibáñez et al., 2014; Quiroz-González et al., 2020), así como por distintas condiciones abióticas como circulación del agua, temperatura, pH, régimen de mareas y sedimento (López et al., 2017, 2023). Por ello, es fundamental realizar más investigaciones enfocadas hacia el estudio de la variación espacio-temporal de los organismos en periodos distintos a un ciclo anual en diferentes localidades de México, con el objetivo de sentar las bases pertinentes para su conservación y posterior manejo ante el incremento del desarrollo de infraestructuras turísticas/urbanas en zonas como Guerrero, que pueden tener un efecto desfavorable sobre las comunidades intermareales a largo plazo (Zamorano y Leyte-Morales, 2009).
Agradecimientos
Al proyecto DGAPA-PAPIIT, UNAM (IN220714), al Registro Nacional de Pesca y Acuacultura por el permiso para la recolección de material biológico (DF00000208). A Norma López por el préstamo de las instalaciones de la UMDI-Zihuatanejo, a Carlos Candelaria por su apoyo técnico en campo y a Isabel Bieler, por su apoyo en la toma de las fotografías de los ejemplares para
este estudio.
Referencias
Abbott, I. A. (1999). Marine red algae of the Hawaiian Island. Honolulu, Hawaii: Bishop Museum Press.
Abbott, I. A. y Hollenberg, G. J. (1976). Marine algae of California. Stanford, California: Sanford University Press.
Aguilar, L. G. (2017). Estructura comunitaria de los moluscos (gasterópodos pateliformes y poliplacóforos) asociados a los ensambles macroalgales en el intermareal rocoso de Ixtapa-Zihuatanejo, Guerrero, México (Tesis de maestría). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Aguilar-Estrada, L. G., Ortigosa, D., Urbano, B. y Reguero, M. (2014). Análisis histórico de los gasterópodos de la laguna arrecifal de Isla Verde, Veracruz, México. Revista Mexicana de Biodiversidad, 85, 502–512. https://doi.org/10.7550/rmb.
33802
Aguilar-Estrada, L. G., Ruiz-Boijseauneau, I. y Rodríguez, D. (2017). Estadios juveniles de las especies de gasterópodos pateliformes y de poliplacóforos (Mollusca) asociados a macroalgas intermareales de Guerrero, México. Revista Mexicana de Biodiversidad, 88, 280–299. https://doi.org/10.1016/j.rmb.2017.03.021
Aguilar-Estrada, L. G., Quiroz-González, N., Ruiz-Boijseauneau, I., Álvarez-Castillo, L. y Rodríguez, D. (2022). Algal epibiont species on Chiton articulatus (Mollusca: Polyplacophora) from a rocky intertidal coast from the Mexican Tropical Pacific. Revista Mexicana de Biodiversidad, 93, e934163. https://doi.org/10.22201/ib.20078706e.2022.93.4163
Aguilera, M. A. (2011). The functional roles of herbivores in the rocky intertidal systems in Chile: A review of food preferences and consumptive effects. Revista Chilena de Historia Natural, 84, 241–261.
Aguillón, A. (2011). Variación espacio-temporal del recluta-
miento en Mollusca y Echinodermata en bahía de La Paz, Baja California Sur, México (Tesis de maestría). Ciudad de México: Centro Interdisciplinario de Ciencias Marinas, Instituto Politécnico Nacional.
Airoldi, L. (2003). The effects of sedimentation on rocky coast assemblages. Oceanography and Marine Biology an Annual Review, 41, 167–171.
Anderson, M., Gorley, R. y Clarke, K. (2008). Permanova+ for PRIMER: guide to software and statistical methods, PRIMER-E, Plymouth.
Bakus, G. J. (2007). Quantitative analysis of marine biological communities: field biology and environmental. Hoboken, New Jersey: John Wiley and Sons Inc.
Baqueiro, E. (1979). Sobre la distribución de Megapitaria aurantiaca (Sowerby), M. squalida (Sowerby) y Dosinia ponderosa (Gray) en relación a la granulometría del sedimento (Bivalvia: Veneridae). Anales del Centro de Ciencias del Mar y Limnología, Universidad Nacional Autónoma de México, 6, 25–32.
Baqueiro, E. y Aldana, D. (2000). A review of reproductive patterns of bivalve mollusks from Mexico. Bulletin of Marine Science, 66, 13–27.
Baqueiro, E. y Aldana, D. (2003). Patrones en la biología poblacional de moluscos de importancia comercial en México. Revista de Biología Tropical, 51, 97–107.
Baqueiro, E. y Masso, J. A. (1988). Variaciones poblacionales y reproducción de dos poblaciones de Chione undatella (Sowerby, 1835) bajo diferentes regímenes de pesca en la bahía de La Paz, BCS México. Ciencia Pesquera, 6, 51–67.
Barba-Marino, F., Flores-Rodríguez, P., Flores-Garza, R., García-Ibáñez, S. y Arana-Salvador, D. G. (2010). Biodiversidad y zonificación de la comunidad de moluscos, que habita el sustrato rocoso en dos sitios con distinta acción del oleaje, en la Isla “La Roqueta”, Acapulco, Guerrero, México. En L. J. Rangel, J. Gamboa, S. L. Arriaga y W. M. Contreras (Eds.), Perspectiva en malacología mexicana (pp. 44–56). Villahermosa: Universidad Juárez Autónoma de Tabasco.
Bartlett, M.S. (1937). Properties of sufficiency and statistical test. Proceedings of the Royal Society A, 160, 268–282. https://doi.org/10.1098/rspa.1937.0109
Baumgartner, T. R. y Christensen, N. (1985). Coupling of the Gulf of California to large-scale interannual climatic variability. Journal of Marine Research, 43, 825–848.
Benedetti-Cecchi, L. (2001). Variability in abundance of algae and invertebrates at different spatial scales on rocky sea shores. Marine Ecology Progress Series, 215, 79–92.
Benedetti-Cecchi, L., Rindi, F., Bertocci, I., Bulleri, F. y Cinelli, F. (2001). Spatial variation in development of epibenthic assemblages in a coastal lagoon. Estuarine, Coastal and
Shelf Science, 52, 659–668. https://doi.org/10.1006/ecss.20
01.0775
Borges, L., Merckelbach, L. M., Sampaio, Í. y Cragg, S. M. (2014). Diversity, environmental requirements, and biogeography of bivalve wood-borers (Teredinidae) in European coastal waters. Frontiers in Zoology, 11, 1–13. https://doi.org/10.1186/1742-9994-11-13
Bouchet, P., Rocroi, J. P., Bieler, R., Carter, J. G. y Coan, E. V. (2010). Nomenclator of bivalve families with a classification of bivalve families. Malacologia, 52, 1–185. https://doi.org/
10.4002/040.052.0201
Cadena-Cárdenas, L., Méndez-Rodríguez, L., Zenteno-Savín, T., García-Hernández, J. y Acosta-Vargas, B. (2009). Heavy metal levels in marine mollusks from areas with, or without, mining activities along the Gulf of California, Mexico. Archives of Environmental Contamination and Toxicology, 57, 96–102. https://doi.org/10.1007/s00244-008-9236-0
Castro-Mondragón, H., Flores-Garza, R., Valdez-González, A., Flores-Rodríguez, P., García-Ibáñez, S. y Rosas-Acevedo, J. L. (2016). Diversidad, especies de mayor importancia y composición de tallas de los moluscos en la pesca ribereña en Acapulco, Guerrero, México. Acta Universitaria, 26, 24–34. https://doi.org/10.15174/au.2016.1025.
Ceccherelli, V. U. y Rossi, R. (1984). Settlement, growth and production of the mussel Mytilus galloprovincialis. Marine Ecology Progress Series, 16, 173–184.
Cerros-Cornelio, J. C., Flores-Garza, R., Landa-Jaime, V., García-Ibáñez, S., Rosas-Guerrero, V., Flores-Rodríguez, P. et al. (2021). Composición de especies e ingreso económico por la pesca ribereña de moluscos en la Costa Grande de Guerrero México. Revista Bio Ciencias, 8, e1054. https://doi.org/10.15741/revbio.08.e1054
Chemello, R. y Milazzo, M. (2002). Effect of algal architecture on associated fauna: some evidence from phytal molluscs. Marine Biology, 140, 981–990. https://doi.org/10.1007/s00
227–002–0777–x
Cisneros, F. (2016). Estructura comunitaria de gasterópodos asociados a macroalgas en el litoral rocoso de Ixtapa-Zihuatanejo, Guerrero, México (Tesis). Facultad de Ciencias,
Universidad Nacional Autónoma de México. Ciudad de México.
Clarke K. R. y Gorley, R. N. (2006) PRIMER v6: User Manual/Tutorial. PRIMER-E Ltd, Plymouth, Reino Unido.
Coan, E. V. (1999). The eastern Pacific species of Sphenia (Bivalvia: Myidae). Nautilus-Sanibel, 113, 103–120. https://doi.org/10.5962/bhl.part.2019
Coan, E. V., Scott, P. V. y Bernard, F. R. (2000). Bivalve seashells of western North America. Santa Bárbara, California: Santa Barbara Museum of Natural History Monographs.
Coan, E. V. y Valentich-Scott, P. (2006). Marine Bivalvia. En C. F. Sturm, T. A. Pearce y A. Valdés (Eds.), The mollusks: a guide to their study, collection, and preservation (pp. 339–349). Pittsburgh, Pennsylvania: American Malacological Society.
Coan E.V. y Valentich-Scott P. (2012). Bivalve seashells of tropical West America. Marine bivalve mollusks from Baja California to Peru. Santa Barbara: Santa Barbara Museum of Natural History Monographs.
Colman, J. (1940). On the faunas inhabiting intertidal seaweeds. Journal of the Marine Biological Association of the United Kingdom, 24, 129–183. https://doi.org/10.1017/S0025315400054503
Dawson, E. Y. (1949). Resultados preliminares de un reco-
nocimiento de las algas marinas de la costa del pacífico de México. Revista de la Sociedad Mexicana de Historia Natural, 9, 215–255.
Dawson, E. Y. (1953). Marine red algae of Pacific Mexico. Part 1. Bangiales to Corallinaceae subf. Corallinoidae. Allan Hancock Pacific Expedition, 17, 1–239.
Dawson, E. Y. (1954). Marine red algae of Pacific Mexico. Part 2. Cryptonemiales (cont.). Allan Hancock Pacific Expedition, 17, 241–397.
Dawson, E. Y. (1960). Marine red algae of Pacific Mexico. Part 3. Cryptonemiales, Corallinaceae subf. Melobesioidae. Pacific Naturalist, 2, 3–125.
Dawson, E. Y. (1961). Marine red algae of Pacific Mexico. Part 4. Gigartinales. Pacific Naturalist, 2, 191–343.
Dawson, E. Y. (1963). Marine red algae of Pacific Mexico part 8. Ceramiales: Dasyaceae, Rhodomelaceae. Nova Hedwigia, 6, 401–481.
Dawson, E. Y. y Beaudette, P. T. (1959). Field notes from the 1959 Eastern Pacific Cruise of the Stella Polaris. Pacific Naturalist, 1, 1–24.
Díaz, A. y Ortlieb, L. (1993). El fenómeno “El Niño” y los moluscos de la costa peruana. Bulletin de l’Institut Français d’Études Andines, 22, 159–177.
Dobretsov, S. V. (1999). Effects of macroalgae and biofilm on settlement of blue mussel (Mytilus edulis l.) larvae, Bio-
fouling, 14, 153–165. https://doi.org/10.1080/089270199093
78406
Esqueda-González, M., Ríos-Jara, E., Galván-Villa, C. M. y Rodríguez-Zaragoza, F. A. (2014). Species composition, richness, and distribution of marine bivalve molluscs in Bahía de Mazatlán, México. Zookeys, 399, 43–69. https://doi.org/10.3897/zookeys.399.6256
Feely, R. A., Sabine, C. L., Lee, K., Berelson, W., Kleypas, J., Fabry, V. J. et al. (2004). Impact of anthropogenic CO2 on the CaCO3 system in the oceans. Science, 305, 362–366. https://doi.org/10.1126/science.1097329
Fitch, J. E. (1953). Common marine bivalves of California. Sacramento, California: California Department of Fish and Game.
Flores, P. (2004). Estructura de la comunidad de moluscos del mesolitoral superior en las playas de facie rocosa del estado de Guerrero, México (Tesis doctoral). Facultad de Ciencias Biológicas, Universidad Autónoma de Nuevo León. San Nicolás de los Garza.
Flores-Garza, R., Flores-Rodríguez, P., García-Ibáñez, S. y Valdés-González, A. (2007). Demografía del caracol Plicopurpura pansa (Neotaenioglossa: Muricidae) y constitución de la comunidad malacológica asociada en Guerrero, México. Revista de Biología Tropical, 55, 867–878.
Flores-Garza, R., García-Ibáñez, S., Flores-Rodríguez, P., Torreblanca-Ramírez, C., Galeana-Rebolledo, L., Valdés-González, A. et al. (2012). Commercially important marine mollusks for human consumption in Acapulco, México. Natural Resources, 3, 11–17. http://dx.doi.org/10.4236/nr.20
12.31003
Flores-Garza, R., López-Rojas, V., Flores-Rodríguez, P. y Ramírez, C. T. (2014). Diversity, distribution, and composition of the Bivalvia class on the rocky intertidal zone of marine priority region 32, Mexico. Open Journal of Ecology, 4, 961–973 https://doi.org/10.4236/oje.2014.415080
Flores-Garza, R., Torreblanca-Ramírez, C., Flores-Rodríguez, P., García-Ibáñez, S. y Galeana-Rebolledo, L. (2010). Riqueza y análisis de la comunidad malacológica en el mesolitoral rocoso de la playa Tlacopanocha, Acapulco, Guerrero. En L. J. Rangel, J. Gamboa, S. L. Arriaga y W. M. Contreras (Eds.), Perspectiva en malacología mexicana (pp. 125–138). Villahermosa: Universidad Juárez Autónoma de Tabasco.
Flores-Garza, R., Torreblanca-Ramírez, C., Flores-Rodríguez, P., García-Ibáñez, S., Galeana-Rebolledo, L., Valdés-González, A. et al. (2011). Mollusc community from a rocky intertidal zone in Acapulco, Mexico. Biodiversity, 12, 144–153. https://doi.org/10.1080/14888386.2011.625520
Flores-Rodríguez, P., Flores-Garza, R., García-Ibáñez, S. y Valdés-González, A. (2003). Riqueza y diversidad de la malacofauna del mesolitoral rocoso de la Isla la Roqueta, Acapulco, Guerrero, México. Ciencia, Revista de Investigación Científica, 11, 5–14.
Flores-Rodríguez, P., Flores-Garza, R., García-Ibáñez, S. y Valdés-González, A. (2007). Variación en la diversidad malacológica del mesolitoral rocoso en Playa Troncones, La Unión, Guerrero, México. Revista Mexicana de Biodiversidad, 78, 33–40. https://doi.org/10.22201/ib.20078706e.2007.002.298
Flores-Rodríguez, P., Flores-Garza, R., García-Ibáñez, S., Valdés-González, A., Violante-González, J., Santiago, C. E. et al. (2012). Mollusk species richness on the rocky shores of the state of Guerrero, Mexico, as affected by rains and their geographical distribution. Natural Resources, 3, 248–260. http://dx.doi.org/10.4236/nr.2012.34032
Forster, S. y Zettler, M. L. (2004). The capacity of the filter-feeding bivalve Mya arenaria L. to affect water transport in sandy beds. Marine Biology, 144, 1183–1189. https://doi.org/10.1007/s00227-003-1278-2
Galeana-Rebolledo, L., Flores-Garza, R., Torreblanca-Ramírez, C., García-Ibáñez, S., Flores-Rodríguez, P. y López-Rojas, V. I. (2012). Biocenosis de Bivalvia y Polyplacophora del intermareal rocoso en playa Tlacopanocha, Acapulco, Guerrero, México. Latin American Journal of Aquatic Research, 40, 943–954. http://dx.doi.org/10.3856/vol40-
issue4-fulltext-11
Galeana-Rebolledo, L., Flores-Garza, R., Violante-González, J., Flores-Rodríguez, P., García-Ibáñez, S., Landa-Jaime, V. et al. (2018). Socioeconomic aspects for coastal mollusk commercial fishing in Costa Chica, Guerrero, México. Natural Resources, 9, 229–241. https://doi.org/10.4236/nr.
2018.96015
García-Domínguez, F. A., Haro-Hernández, D., García-Cuellar, Á., Villalejo-Fuerte, M. y Rodríguez-Astudillo, S. (2008). Ciclo reproductivo de Anadara tuberculosa (Sowerby, 1833) (Arcidae) en Bahía Magdalena, México. Revista de Biología Marina y Oceanografía, 43, 143–152. http://dx.doi.org/10.4067/S0718-19572008000100015.
Gama, N. (2019). Aspectos ecológicos de la comunidad de gasterópodos (Mollusca: Gastropoda) asociados a ensambles macroalgales en el intermareal rocoso de Ixtapa Zihuatanejo, Guerrero, México (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Gama-Kwick, N., Aguilar-Estrada, L. G., Quiroz-González, N. y Ruiz-Boijseuneau, I. (2021). Nuevos registros de gasterópodos (Mollusca) asociados a macroalgas inter-
mareales de Guerrero, México. Revista Mexicana de Biodiversidad, 92, 1–18. http://dx.doi.org/10.22201/ib.200
78706e.2021.92.3441
Garcés, J. L. (2011). Micromoluscos bivalvos de la Bahía de Acapulco Guerrero, México: composición específica y diversidad (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
García, M. (2009). Estructura comunitaria de la fauna asociada a algas submareales en tres sitios contrastantes en Zihuatanejo, Guerrero (Tesis de maestría). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
García-Cubas, A. (1981). Moluscos de un sistema lagunar tropical en el sur del Golfo de México (Laguna de Términos, Campeche). Publicaciones especiales-Instituto de Ciencias del Mar y Limnología, 5, 1–379.
García-Cubas, A. y Reguero, M. (2007). Catálogo ilustrado de moluscos bivalvos del Golfo de México y Mar Caribe. Ciudad de México: Instituto de Ciencias del Mar y Limnología, Universidad Nacional Autónoma de México.
García-Ibáñez, S., Flores-Rodríguez, P., Navarro, J. T. N., Garza, R. F. y Moreno, I. E. B. (2014). Respuesta del carnívoro Plicopurpura pansa (Mollusca: Gastropoda) y el herbívoro Chiton articulatus (Mollusca: Polyplacophora) a factores ambientales en Acapulco, México. CienciaUAT, 8, 11–21.
García-Robledo, E., Corzo, A., van Bergeijk, S. y Yúfera, M. (2008). Impacto de las acumulaciones de macroalgas en la comunidad biológica intermareal. Revista de la Sociedad Gaditana de Historia Natural, 8, 115–137.
Gibbons, M. J. (1988). The impact of sediment accumulation, relative habitat complexity and elevation on rocky shore meiofauna. Journal of Experimental Marine Biology Ecology, 122, 225–241. https://doi.org/10.1016/0022-0981(88)90125-6
Gosling, E. (2015). Marine bivalve molluscs. Chichester, West Sussex: John Wiley y Sons.
Guiry, M. D. y Guiry, G. M. (2024). AlgaeBase. National University of Ireland, Galway. Recuperado el 12 enero, 2024 de: http://www.algaebase.org
Guzmán, A. R. (2022). Estructura comunitaria de bivalvos asociados a macroalgas en el litoral rocoso de Muelle Municipal, Zihuatanejo, Guerrero, México (Tesis). Facultad de Ciencias. Universidad Nacional Autónoma de México. Ciudad de México
Hernández-Moreno, E. P., Romo-Piñera, A. K., Fernández-Rivera Melo, F. J., Aguilar-Cruz, C. A., Reyes-Bonilla, H. y López-Vivas, J. M. (2020). Reproductive Study of Megapitaria aurantiaca (Sowerby, 1831) (Bivalvia: Veneridae) in Puerto Libertad, Sonora, Mexico. Journal of Shellfish Research, 39, 441–447.
Holguín-Quiñones, O. E. y González-Pedraza, A. C. (1989). Moluscos de la franja costera del estado de Oaxaca, México. México D.F.: Dirección de Bibliotecas y Publicaciones, Instituto Politécnico Nacional.
Horton, T., Kroh, A., Ahyong, S., Bailly, N., Boyko, C. B. y Brandão, S. N. (2024. World Register of Marine Species. Recuperado el 14 de marzo, 2024: http://www.marinespecies.org
Huovinen, P., Gómez, I. y Lovengreen, C. (2006). A five-year study of solar ultraviolet radiation in Southern Chile (39° S): Potential impact on physiology of coastal marine algae? Photochemistry and Photobiology, 82, 515–522. https://doi.org/10.1562/2005-07-05-RA-601
IMTA (Instituto Mexicano de Tecnología del Agua). (2010). Estudio de clasificación de la Bahía de Ixtapa-Zihuata-
nejo. No. FON-CNA-2004-02-016. Informe final. Comisión Nacional del Agua.
Jiménez-Valverde, A. y Hortal, J. (2003). Las curvas de acumulación de especies y la necesidad de evaluar la calidad de los inventarios biológicos. Revista Ibérica de Aracnología, 8, 151–161.
Jorgensen, C. B. (1996). Bivalve filter feeding revisited. Marine Ecology Progress Series, 142, 287–302.
Jover-Capote, A. y Diez, Y. L. (2017). Abundancia de moluscos en mantos de macroalgas del mesolitoral rocoso en la costa suroriental de Cuba. Amici Molluscarum, 25, 27–43.
Keen, A. M. (1971). Sea shells of Tropical West America Marine mollusks from Baja California to Peru. Standford, California: Stanford University Press.
Kuk-Dzul, J. G., Padilla, J. G., Torreblanca, C., Flores, R., Flores, P. y Muñiz, X. I. (2019). Structure of molluscan communities in shallow subtidal rocky bottoms of Acapulco, Mexico. Turkish Journal of Zoology, 43, 465–479.
Landa-Jaime, V. y Arciniega-Flores, J. (1998). Macromoluscos bentónicos de fondos blandos de la plataforma continental de Jalisco y Colima, México. Ciencias Marinas, 24, 155–l67.
Landa-Jaime, V., Michel-Morfín, E., Arciniega-Flores, J., Castillo-Vargasmachuca, S. y Saucedo-Lozano, M. (2013). Moluscos asociados al arrecife coralino de Tenacatita, Jalisco, en el Pacífico central mexicano. Revista Mexicana de Biodiversidad, 84, 1121–1136. https://doi.org/10.7550/rmb.32994
Lee, R. E. (2008). Basic characteristics of the algae. En R. E. Lee (Ed.), Phycology (pp. 3-29). Nueva York: Cambridge University Press.
Lesser, H. (1984). Prospección sistemática y ecológica de los moluscos bentónicos de la plataforma continental del estado de Guerrero, México (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Levene, H. (1960) Robust tests for equality of variances. En I. Olkin (Ed.), Contributions to probability and statistics (278–292). Palo Alto, California: Standford University Press.
López, N. A., (1993). Caracterización de la ficoflora sublitoral de Acapulco y Zihuatanejo, Gro. (Tesis). Facultad de Cien-
cias, Universidad Nacional Autónoma de México. Ciudad de México.
López, N., Candelaria, C. y Ramírez-García, P. (2023). Assessment of macroalgae coverage in a scarcely studied deep rocky reef in the tropical eastern Mexican Pacific. Latin American Journal of Aquatic Research, 51, 23–33. https://doi.org/10.3856/vol51-issue1-fulltext-2920
López, N., Candelaria, C., Ramírez-García, P. y Rodríguez, D. (2017). The structure of tropical turf-forming algae assemblages. Zihuatanejo Bay, México. Latin American Journal of Aquatic Research, 45, 329–340 http://dx.doi.org/10.3856/vol45-issue2-fulltext-9
López-Rojas, V. I., Flores-Garza, R., Flores-Rodríguez, P., Torreblanca-Ramírez, C. y García-Ibáñez, S. (2017). La clase Bivalvia en sitios rocosos de las Regiones Marinas Prioritarias en Guerrero, México: riqueza de especies, abundancia y distribución. Hidrobiológica, 27, 69–86.
Lozada, O. (2010). Actualización sistemática de los bivalvos de la Colección Malacológica Dr. Antonio García-Cubas del Instituto de Ciencias del Mar y Limnología (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Lozada, O. (2015). Presencia de metales pesados en Isogno-
mon alatus (Gmelin, 1791) de la laguna Tampamachoco, Veracruz (Tesis de maestría). Facultad de Ciencias Bio-
lógicas y Agropecuarias, Universidad Veracruzana. Tuxpan, Veracruz.
Magurran, A. E. (2004). Measuring biological diversity. Malden, Massachusetts: Blackwell Publishing.
Morales, R., Vélez, H., Mejía, A. Ramírez, I., Izurierta, J. y Saldaña, P. (2008). Hidrodinámica de la Bahía de Zihuata-
nejo. XXIII Congreso Latinoamericano de Hidráulica, 2 al 6 de septiembre de 2008, Cartagena de Indias, Colombia.
Moreno, C. A. (1995). Macroalgae as a refuge from predation for recruits of the mussel Choromytilus chorus (Molina, 1782) in southern Chile. Journal of Experimental Marine Biology and Ecology, 191, 181–193.
Moreno, C. E. (2001). Métodos para medir la biodiversidad. Zaragoza, España: M&T–Manuales y Tesis SEA.
Oksanen, J., Guillaume-Blanchet, F., Friendly, M., Kindt, R. Legendre, P., McGlinn, D. et al. (2019). Vegan: community ecology package. R package versión 2.5-6. https://CRAN.R-project.org/package=vegan
Olabarria, C. y Chapman, M. G. (2001). Comparison of patterns of spatial variation of microgastropods between two contrasting intertidal habitats. Marine Ecology Progress Series, 220, 201–211. https://doi.org/doi:10.3354/meps220201
Olafsson, E. B. (1986). Density dependence in suspension-feeding and deposit-feeding populations of the bivalve Macoma balthica: a field experiment. Journal of Animal Ecology, 55, 517–526. https://doi.org/10.2307/4735
Oyarzún, P. A., Toro, J. E., Jaramillo, R., Guiñez, R., Briones, C. y Astorga, M. (2011). Ciclo gonadal del chorito Mytilus chilensis (Bivalvia: Mytilidae) en dos localidades del sur de Chile. Latin American Journal of Aquatic Research, 39, 512–525.
Paredes, C., Tarazona, J., Canahuire, E., Romero, L., Cornejo, O. y Cardoso, F. (1998). Presencia de moluscos tropicales de la provincia panameña en la costa central del Perú y su relación con los eventos “El Niño”. Revista Peruana de Biología, 5, 123–128. https://doi.org/10.15381/rpb.v5i2.8330
Pérez, L. M. (2013). Descripción espacio-temporal de la temperatura superficial del mar en el Pacífico Sur Mexicano de 1996-2009 (Tesis). Universidad del Mar. Puerto Ángel, Oaxaca.
Prathep, A., Marrs, R. H. y Norton, T. A. (2003). Spatial and temporal variations in sediment accumulation in an algal turf and their impact on associated fauna. Marine Biology, 142, 381–390. https://doi.org/10.1007/s00227–002–0940–4
Quirós-Rodríguez, J. y Campos, N. H. (2013). Moluscos asociados a ensamblajes macroalgales en el litoral rocoso de Córdoba, Caribe colombiano. Boletín de Investigaciones Marinas y Costeras, 42, 101–120.
Quiroz-González, N., Aguilar-Estrada, L. G., Ruiz-Boijseauneau, I. y Rodríguez, D. (2020). Biodiversidad de algas epizoicas en el Pacífico tropical mexicano. Acta Botanica Mexicana, 127, 1-22. https://doi.org/10.21829/abm127.2020.1645
R Core Team (2020). R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. https://www.R-project.org/
Reguero, M., y García-Cubas, A. (1989). Moluscos de la plataforma continental de Nayarit: sistemática y ecología (cuatro campañas oceanográficas). Anales del Instituto de Ciencias del Mar y Limnología, 16, 33–58.
Renaud, P. E., Riggs, S. R., Ambrose Jr, W. G., Schmid, K. y Snyder, S. W. (1997). Biological-geological interactions: storm effects on macroalgal communities mediated by sediment characteristics and distribution. Continental Shelf Research, 17, 37–56. https://doi.org/10.1016/0278-43
43(96)00019-2
Ríos-Jara, E., López-Uriarte, E., Pérez-Peña, M. y Juárez-Carrillo, E. (2003). Nuevos registros de escafópodos para las costas de Jalisco y Colima, México. Hidrobiológica, 13, 167–170.
Rodríguez, D., López, N. y González-González, J. (2008). Gelidiales (Rhodophyta) en las costas del Pacífico mexicano con énfasis en las especies tropicales. En A. Sentíes y K. M. Dreckmann (Eds.), Monografías ficológicas, Vol. 3 (pp. 27–74). Ciudad de México: Universidad Autónoma Metropolitana.
Román-Contreras, R., Cruz-Abrego, F. M. e Ibáñez-Aguirre, A. L. (1991). Observaciones ecológicas de los moluscos de la zona intermareal rocosa de la Bahía de Chamela, Jalisco, México. Anales del Instituto de Biología, Universidad Nacional Autónoma de México, Serie Zoología, 62, 17–31.
Rosenberg, R. (1977). Effects of dredging operations on estuarine benthic macrofauna. Marine Ecology Progress Series, 62, 185–202.
Salazar-Vallejo, S. I. y González, E. (1990). Ecología costera en la región de La Mancha, Veracruz. La Ciencia y el Hombre, 6,101–120.
Salcedo-Martínez, S., Green, G., Gamboa-Contreras, A. y Gómez,
P. (1988). Inventario de macroalgas y macroinvertebrados bénticos, presentes en áreas rocosas de la región de Zihuatanejo, Guerrero, México. Anales del Instituto de Ciencias del Mar y Limnología, Universidad Nacional Autónoma de México, 15, 73–95.
Sánchez, M. (2014). Base de datos de los moluscos presentes en la colección biológica de la Secretaría de Marina-Armada de México (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Santos, L., Souza, J., Lima, S. y Guimarães, C. (2020). Diversity of bivalve mollusks associated with macroalgae on the continental shelf in the states of Alagoas, Sergipe and Bahia, Northeastern Brazil. Zoological Studies, 59, e58. https://doi.org/10.6620/ZS.2020.59-58
Schulien, J. A., Adams, J. y Felis, J. (2020). Pacific continental shelf environmental assessment (PaCSEA): characterization of seasonal water masses within the Northern California current system using airborne remote sensing off Northern California, Oregon, and Washington, 2011–2012. Camarillo, California: US Department of the Interior, Bureau of Ocean Energy Management/ Pacific OCS Region/ OCS Study BOEM.
Seed, R. (1969a). The ecology of Mytilus edulis L. (Lamellibranchiata) on exposed rocky shores. Oecologia, 3, 277–316. https://doi.org/10.1007/BF00390380
Seed, R. (1969b). The ecology of Mytilus edulis L. (Lamelli-
branchiata) on exposed rocky shores: II. Growth and mortality. Oecologia, 3, 317–350. https://www.jstor.org/stable
/4214550
Shaw, W. N., Hassler, T. J. y Moran, D. P. (1988). California sea mussel and bay mussel: species profiles. Life histories and environmental requirements of coastal fishes and invertebrates (Pacific Southwest). Washington D.C.: Fish and Wildlife Service.
Sibaja-Cordero, J. A. y Vargas-Zamora, J. A. (2006). Zonación vertical de epifauna y algas en litorales rocosos del Golfo de Nicoya, Costa Rica. Revista de Biología Tropical, 54, 49–67.
Siegel, S. (1990). Estadística no paramétrica: aplicada a las ciencias de la conducta. Ciudad de México: Trillas.
Stella, C., Vijayalakshmi, S. y Siva, J. (2010). Two new records of bivalve species of Mytilidae family from Palk Bay and Mandapam area-south-east coast of India. Global Journal of Environmental Research, 4, 40–42.
Steneck, R. S. y Dethier, M. N. (1994). A functional group approach to the structure of algal-dominated communities. Oikos, 69, 476–498. https://doi.org/10.2307/3545860
Steneck, R. S. y Watling, L. (1982). Feeding capabilities and limitation of herbivorous molluscs: a functional group approach. Marine Biology, 68, 299–319. https://doi.org/10.
1007/BF00409596
Suchanek, T. H. (1978). The ecology of Mytilus edulis L. in exposed rocky intertidal communities. Journal of Experimental Marine Biology and Ecology, 31, 105–120. https://doi.org/10.1016/0022-0981(78)90139-9
Taylor, W. R. (1945). Pacific marine algae of the Allan Hancock Expeditions to the Galapagos Islands. Allan Hancock Pacific Expeditions, 12, 1–528.
Torreblanca, C. (2010). Análisis de la diversidad y estructura de la comunidad de moluscos del mesolitoral rocoso de Acapulco, Guerrero (Tesis).Unidad Académica de Ecología Marina, Universidad Autónoma de Guerrero. Acapulco de Juárez.
Torreblanca-Ramírez, C., Flores-Garza, R., Flores-Rodríguez, P., García-Ibáñez, S. y Galeana-Rebolledo, L. (2012). Riqueza, composición y diversidad de la comunidad de moluscos asociada al sustrato rocoso intermareal de playa Parque de la Reina, Acapulco, México. Revista de Biología Marina y Oceanografía, 47, 283–294. http://dx.doi.org/10.4067/S0718-19572012000200010
Torroglosa, M. E. (2015). Biología reproductiva y crecimiento de Brachidontes rodriguezii (d´Orbign 1846) en sustratos duros artificiales en playas arenosas de la provincia de Buenos Aires (Tesis doctoral). Facultad de Ciencias Exactas y Naturales, Universidad de Buenos Aires. Argentina.
Trasviña, A. y Andrade, C. A. (2002). La circulación costera del Pacifico Tropical Oriental, con énfasis en la Alberca Cálida Mexicana (ACM). En Corcas (Eds.), Circulación oceánica y climatología tropical en México y Colombia (pp. 9–37). Bogotá: Consejo Nacional de Acreditación-Ministerio de Educación.
UNAM (Universidad Nacional Autónoma de México) (2013). Laguna Las Salinas Zihuatanejo. Proyecto ejecutivo UNAM 1989-1994 basado en plano base FONATUR-SCT. Universidad Nacional Autónoma de México, México. Recuperado el 14 enero, 2024 de https://mmacreactive.wordpress.com/
Urbano, B. (2004). Estructura comunitaria de gasterópodos de Zihuatanejo, Guerrero (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Valdés-González, A., Flores-Rodríguez, P., Flores-Garza, R. y García-Ibáñez, S. (2004). Molluscan communities of the rocky intertidal zone at two sites with different wave action on Isla La Roqueta, Acapulco, Guerrero, Mexico. Journal of Shellfish Research, 23, 875–880.
Vassallo, A., Dávila, Y., Luviano, N., Deneb-Amozurrutia, S., Vital, X. G., Conejeros, C. A. et al. (2014). Inventario de invertebrados de la zona rocosa intermareal de Montepío, Veracruz, México. Revista Mexicana de Biodiversidad, 85, 349–362. https://doi.org/10.7550/rmb.42628
Vázquez, P. (2009). Efecto del sedimento retenido en la estructura de los ensambles algales submareales (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Vega, C., Olabarria, C. y Carballo, J. L. (2008). Variación espacio-temporal de moluscos y macroalgas en sustratos rocosos intermareales en la bahía de Mazatlán. Ciencia y Mar, 34, 3–16.
Velez, J. y Zeballos, J. (1985). Ampliación de la distribución de algunos peces e invertebrados durante el fenómeno “El Niño” 1982–1983. En W. Arntz, A. Landa y J. Tarazona (Eds.), “El Niño” su impacto en la fauna marina (pp. 173– 180). Callao, Perú: Boletín del Instituto del Mar del Perú.
Villalpando, E. (1986). Diversidad y zonación de moluscos de facie rocosa Isla Roqueta, Acapulco, Gro. (Tesis). Facultad de Ciencias, Universidad Nacional Autónoma de México. Ciudad de México.
Villegas-Maldonado, S., Neri-García, E., Flores-Garza, R., García-Ibáñez, S., Flores Rodríguez, P. y Arana-Salvador, D. G. (2007). Datos preliminares de la diversidad de moluscos para el consumo humano que se expenden en Acapulco, Guerrero. En E. Ríos-Jara, M. C. Esqueda-González y C. M Galván-Villa (Eds.), Estudios sobre la malacología y conquiliología en México (pp. 57-59). Guadalajara: Universidad de Guadalajara.
Ward, J. E., Sanford, L. P., Newell, R. I. E. y MacDonald, B. A. (1998). A new explanation of particle capture in suspension-feeding bivalve molluscs. Limnology and Oceanography, 43, 741–752.
Wyrtki, K. 1966. Oceanography of the eastern equatorial Pacific Ocean. Oceanography and marine biology: an annual review, 4, 33–68.
Yang, J. L., Satuito, C. G., Bao, W. Y. y Kitamura, H. (2007). Larval settlement and metamorphosis of the mussel Mytilus galloprovincialis on different macroalgae. Marine Biology, 152, 1121–1132. https://doi.org/10.1007/s00227-007-0759-0
Zamorano, P. y Leyte-Morales, G. E. (2009). Equinodermos asociados a formaciones arrecifales en Zihuatanejo y Acapulco, Guerrero, México. Boletín de Investigaciones Marinas y Costeras-INVEMAR, 38, 7–28.
Decreased movements of adult female mule deer during winter in arid southwestern North America
Louis C. Bender a, Jon C. Boren b, Shad Cox c, Erik Joaquín Torres-Romero d, e, *
a New Mexico State University, Department of Extension Animal Sciences and Natural Resources, PO Box 30003 MSC 3AE, Las Cruces, New Mexico 88003, USA
b New Mexico State University, Cooperative Extension Service, PO Box 30003 MSC 3AE, Las Cruces, New Mexico 88003, USA
c New Mexico State University, Corona Range and Livestock Research Center, PO Box 392, Corona, New Mexico 88318, USA
d Universidad Politécnica de Puebla, Ingeniería en Biotecnología, Tercer Carril del Ejido, Serrano s/n, San Mateo Cuanalá, Juan C. Bonilla, 72640 Puebla, Mexico
e Tecnológico Nacional de México campus Zacapoaxtla, Subdirección de Investigación y Posgrado, División de Biología, Carretera Acuaco-Zacapoaxtla Km. 8, Col. Totoltepec, 73680 Zacapoaxtla, Puebla, Mexico
*Corresponding author: ejtr23@hotmail.com (E.J. Torres-Romero)
Received: 9 April 2024; accepted: 26 August 2024
Abstract
Deer in northern temperate environments show behavioral and physiological adaptations to conserve energy during winter, including decreased movements. Whether these behaviors persist in warmer temperate environments such as the arid Southwest has received little consideration. We compared daily movements as estimated by continuous-time movement models and minimum subdaily (4 h) straight-line movements of adult female mule deer between winter and spring-autumn seasons in south-central New Mexico. Deer moved less during winter daily (2.90 vs. 4.34 km/d) and subdaily (302 vs. 409 m). Similarly, for deer for which movement data for successive seasons were available, movements between successive seasons were less during the winter (daily = -1.05 km/d; subdaily = -91 m) than the following or preceding spring-autumn. Our results support conservation of decreased movements during winter in the less extreme winters of the arid Southwest. Because some proximate stimuli (i.e., deep snow, very cold temperatures) associated with energy conservation behaviors are lacking in the arid Southwest, our results further support low forage quality and availability being the primary drivers of this behavior.
Keywords: Energy conservation; Movements; Mule deer; New Mexico
© 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/).
Disminución de los movimientos de las hembras adultas de venado bura durante el invierno en el árido suroeste de América del Norte
Resumen
Los ciervos en entornos templados del norte muestran adaptaciones de comportamiento y fisiológicas para conservar energía durante el invierno, incluyendo una disminución en sus movimientos. Se ha explorado poco si estos comportamientos persisten en ambientes templados más cálidos, como el suroeste árido. Comparamos los movimientos diarios, mediante modelos de movimiento continuo y movimientos mínimos en línea recta subdiarios (4 horas) de hembras adultas de venado bura entre las estaciones de invierno y primavera-otoño en el centro-sur de Nuevo México. Los ciervos se movieron menos durante el invierno, tanto diario (2.90 vs. 4.34 km/día) como subdiario (302 vs. 409 m). Además, para ciervos con datos de movimiento en estaciones sucesivas, los movimientos en invierno fueron menores (diarios = -1.05 km/día, subdiarios = -91 m) en comparación con la primavera-otoño previa o siguiente. Nuestros resultados respaldan la disminución de los movimientos durante el invierno en los inviernos menos extremos del suroeste árido. Dado que algunos estímulos inmediatos (por ejemplo, nieve profunda, temperaturas muy frías) asociados con comportamientos de conservación de energía están ausentes en el suroeste árido, es evidente que nuestros resultados apoyan que la baja calidad y disponibilidad de forraje son los principales factores que impulsan este comportamiento.
Palabras clave: Conservación de energía; Movimientos; Venado bura; Nuevo México
Introduction
Deer in northern temperate environments of North America employ a complex energy conservation strategy during winter, incorporating multiple behavioral and physiological adaptations, including reducing activity (i.e., movements) and limiting feeding while relying on endogenous reserves (Alldredge et al., 1974; Short, 1981; Verme & Ullrey, 1984); growing a highly insulative pelage (Jacobsen, 1980); limiting vascular circulation to the extremities (Parker & Robbins, 1984); and lowering metabolic rate to slow the rate of loss of body reserves (Short, 1981; Silver et al., 1971; Verme & Ullrey, 1984). This strategy is considered primarily an adaptation to conserve energy in response to low forage availability and quality, as well as increased costs of movement associated with snow (Short, 1981; Verme & Ullrey, 1984). Conserving energy by minimizing radiant and convective heat loss was also believed to drive other behaviors such as yarding under dense conifer forest canopies (Marchinton & Hirth, 1984). While initially thought to be a response to extreme cold (Marchinton & Hirth, 1984), yarding likely relates to decreased costs of movement because of reduced snow depths as the presumed temperature-moderating influence of forest canopy (i.e., thermal cover) has been shown to have no real effect on deer condition (Cook et al., 1998; Freddy, 1984). Many of the behavioral aspects (at least) of the winter energy conservation strategy are not invariant, however, and can be affected by proximate stimuli. For example, both movements and feeding periods are reduced less if winter conditions are less severe (Bartmann & Bowden, 1984; Verme & Ullrey, 1984).
Whether these energy conservation behaviors persist in warmer temperate ranges such as the arid southwestern USA and Mexico has received little consideration. In the arid Southwest, deer similarly experience winter seasonality in terms of temperature and precipitation differences (Krausman et al., 1990; McKinney, 2003; Marshal et al., 2008), which affects forage availability and quality (Kemp, 1983; Krausman et al., 1990; McKinney, 2003; Short, 1981). Consequently, mule deer lose most of their endogenous reserves over winter (Bender et al., 2012; Bender & Hoenes, 2017). However, snow is relatively rare and short-lived in much of the arid Southwest, and winter temperatures are higher than in northern environments (Table 1). Therefore, aside from decreased forage availability and quality, many of the potential proximate stimuli (i.e., deep snow, very cold temperatures) associated with energy conservation behaviors are lacking in the arid Southwest. Moreover, because winter is less extreme in terms of minimum temperatures and particularly snowfall, availability of forage may also be less limiting, although forage quality constraints are similarly severe (Bender, 2020; Kemp, 1983; Krausman et al., 1990; McKinney et al., 2003). Consequently, behavioral responses associated with the winter energy conservation strategy may be less pronounced or absent in the arid Southwest.
Table 1
Long-term range of monthly mean high and low temperatures (oC) and monthly snowfall (cm) during Dec.-Feb. and Mar.-Nov. at the Corona Range and Livestock Research Center (CRLRC), Corona, New Mexico USA, and the Cusino Wildlife Research Station, Shingleton, Michigan USA. Cusino was selected as a comparison because of the volume and depth of deer nutritional, physiological, and behavioral research conducted there (Verme & Ullrey, 1984). Also presented are the range of conditions on the CRLRC study area for 2005-2008 study period.
Months | Climatic variable | CRLRC | Cusino | Study |
Dec.-Feb. | Mean high temperatures | 6.7-8.9 | -2.8- -1.7 | 6.4-13.0 |
Mean low temperatures | -5.6- -4.4 | -11.7- -8.3 | -6.3- -1.9 | |
Mean snow accumulations | 13-23 | 69-109 | 7.3-17.1 | |
Mar.-Nov. | Mean high temperatures | 11.7-28.3 | 1.7-22.8 | 14.2-29.1 |
Mean low temperatures | -1.1-13.3 | -6.7-13.3 | -0.5-13.8 | |
Mean snow accumulations | 0-13 | 0-48 | 0-5.1 |
If energy conservation behaviors are maintained in the arid Southwest, mule deer (Odocoileus hemionus) should move less during the winter, conserving body reserves in the face of lower quality and less abundant forage even if the impacts of winter weather are less severe on forage availability and costs of movement. Thus, our goal was to contrast short-term movements between winter and spring-autumn seasons for adult female mule deer in a Chihuahuan desert-short grass prairie habitat in New Mexico, USA to determine whether deer reduce movements during winter as predicted by the winter energy conservation strategy. Specifically, we compared minimum daily and subdaily movement distances of adult female mule deer between winter and spring-autumn.
Materials and methods
Our study was conducted on the Corona Range and Livestock Research Center (CRLRC; 34°15’36” N, 105°24’36” W), an 11,290-ha ranch owned and operated by New Mexico State University and located approximately 22.5 km east of Corona, New Mexico (Fig. 1). CRLRC has an average elevation of 1,900 m asl; mean annual precipitation is 40 cm, 87% of which occurs in the Mar.-Nov. period. Snowfall totals < 74 cm annually. Climate of the CRLRC shows distinct seasonality, although the magnitude of seasonal differences in winter is less than seen in northern temperate deer habitats (Table 1).
Topography of the CRLRC is mostly rolling. Vegetation includes perennial grassland, with scattered sparse to dense pinyon (Pinus edulis) and one-seed juniper (Juniperus monosperma) woodlands and a few shrublands. Free water was abundant and comparably available in both winter and spring-autumn seasons because of numerous permanent water developments, ≥ 1 of which were present within or adjacent to annual home ranges of study deer (Fig. 1). Deer on the CRLRC do not migrate between distinct summer and winter ranges.
We captured and collared ≥ 2.5-year-old female mule deer with GPS/VHF radio-collars (Advanced Telemetry Solution, Asanti, Minnesota, USA) programmed to record a position fix every 4 h, early-December 2005-2007, and April, 2006-2007, as part of a larger study of mule deer ecology including other VHF-only radio-collared individuals (Bender et al., 2011, 2013). Deer were captured using a helicopter by aerial net-gunning or darting with 1.5-1.8 mg of carfentanil citrate and 50-75 mg of xylazine hydrochloride per deer. We aged deer as yearling or adult by tooth wear and replacement (Robinette et al., 1957), determined lactation status (Bender et al., 2011), and treated deer with antibiotics, vitamin E/selenium, vitamin B, and an 8-way Clostridium bacterin to help alleviate capture stress. Following processing, immobilants were antagonized with naltrexone and tolazoline.
We defined seasons as winter = Dec-Feb and spring-autumn = Mar-Nov. These seasons corresponded with both typical seasonal and phenological patterns on the CRLRC and the Chihuahuan desert-short grass prairie habitats of the arid Southwest in general, as well as important periods in the annual cycle of female mule deer in the arid Southwest (Bender et al., 2011, 2012). We estimated daily movements using continuous-time movement modeling (see Fleming et al., 2014, 2016), using the Speed/Distance analysis in ctmmweb (https://ctmm.shinyapps.io/ctmmweb/) (Calabrese et al., 2016, 2021). We used only locations with 3D fixes and DOP < 2, as these had an accuracy of < 3 m in our study area. We also determined minimum subdaily movements, defined as the straight-line distance moved between successive 4 h locations, and calculated seasonal means for each deer.
Figure 1. Topographic hillshade showing locations of annual home ranges of adult female mule deer (Bender et al., 2013), and locations of permanent water sources on the Corona Range and Livestock Research Center (CRLRC), east-central New
Mexico, USA.
We compared movement distances (km from continuous-time movement models [ctmms]; m between successive subdaily locations) between seasons using PROC GLIMMIX in SAS 9.4 (SAS, 1988), using individual deer as a random effect. We also compared mean daily and subdaily movements between successive seasons for individual deer for which we had both winter and subsequent spring-autumn data, or spring-autumn and subsequent winter data, available. We determined mean seasonal differences in movement distances for each successive time period for each individual, and used bootstrapping with N = 1,000 iterations to determine the probability that mean movement distances differed seasonally for this subset of data (see Efron & Tibshirani, 1993).
Additionally, because lactating females enter winter in poorer condition than do dry females and condition subsequently converges between the 2 classes over winter (including on the CRLRC; Bender & Hoenes, 2017), lactation status might affect desire or need to forage and thus movements of deer. However, because deer condition was very low on CRLRC during our study (i.e., lactating females were able to accrue only ≤ 5.7% percent body fat annually at the annual peak in late autumn; Bender et al., 2011, 2013), our GPS/VHF collared sample contained ≤ 2 lactating females each year, and thus we were unable to meaningfully include lactation status in our analyses. Consequently, we explored any potential effect of lactation status on movements by determining the percentile movement distances of lactating females relative to the frequency distribution of dry female movement distances to determine whether lactating females were closer to the mean or extremes of the range of dry females.
Lastly, because our GPS collared sample comprised a limited proportion of the overall radio-collared sample, we compared annual, spring-autumn, and winter home range sizes (Bender et al. 2013; N = 18-27 collared females annually) to see whether movements of our GPS sample were representative of radio-collared deer in general, comprised of N = 18-27 VHF-only collared adult females for each season (Bender et al., 2011, 2013). For this we compared annual and seasonal home range sizes of the 2 classes (GPS/VHF and VHF-only) using PROC GLM (SAS, 1988), specifically testing the year × class interaction. We visually located all deer (i.e., both GPS/VHF and VHF-only) via ground tracking of VHF signals a minimum of once per week with additional location emphasis on spring-autumn locations, and mapped locations using the Geographic Information System software package ArcGIS 10.0 (Environmental Systems Research Institute, Redlands, California, USA) (Bender et al., 2011, 2013). We constructed 95% minimum convex polygon (MCP) annual and seasonal home ranges after determining the minimum number of locations to adequately estimate seasonal home range size by plotting size as a function of number of locations (Bender et al. 2013; Kie et al. 1996). For this comparison we used only VHF visual locations of GPS/VHF collared deer so that both GPS/VHF and VHF-only samples were comprised of comparable data.
Results
We collected GPS movement data for 37 seasonal ranges (Table 2), winter 2005-6 through winter 2007-8, from 6-10 GPS/VHF collared adult females annually (mean = 336 and 472 locations per deer for winter and spring-autumn ranges, respectively). For all deer, the OUF-anisotropic movement model provided the best fit (i.e., AICc < 2 vs. all other models) of deer movements. The OUF-anisotropic movement model is the most general of ctmms, and includes a home range, correlated positions, correlated velocities, and movements varying by direction (Calabrese et al., 2021; Fleming et al., 2014).
Table 2
Mean distance moved by adult female mule deer during winter (Dec.-Feb.) and spring-autumn (Mar.-Nov.) seasons as estimated by continuous-time movement modeling (daily) and subdaily straight line movements between successive 4 h locations (subdaily) on the Corona Range and Livestock Research Center, east-central New Mexico, 2005-2007.
Daily (km) | Subdaily (m) | |||||
Season | Distance | SE | N | Distance | SE | N |
Winter | 2.90 | 0.24 | 24 | 302 | 19 | 24 |
Spring-Autumn | 4.34 | 0.43 | 13 | 409 | 44 | 13 |
Deer moved less during the winter (Table 2) for both daily ctmms (F1,20 = 6.0; p = 0.024) and minimum subdaily straight-line distances (F1,20 = 4.8; p = 0.041); in both cases the magnitude of differences varied among individual deer (t15 < -2.72; p < 0.016). Similarly, for deer for which movement data for successive seasons were available, movements between successive seasons were always less during the winter season (p (winter < spring-autumn) = 1.000) than the following or preceding spring-autumn season (Table 3).
Additionally, movement distances of lactating females (mean = 42nd percentile; range = 34-55th percentile) were always closer to the average than the extremes of the frequency distribution of movement distances of dry females for each period. Last, neither annual, spring-autumn, or winter home MCP range sizes (N = 7-16 per period) differed between GPS/VHF and VHF-only collared deer (F5,39 ≤ 0.94; p ≥ 0.599).
Discussion
Despite much less snow and warmer winter temperatures, mule deer females moved less during winter regardless of movement period (i.e., daily, subdaily), supporting the maintenance of this energy conservation behavior in mule deer in the arid Southwest (Verme & Ullrey, 1984). Because deep snow cover and very cold temperatures are typically lacking in the arid Southwest, our results further support low forage quality and availability being the primary drivers of this behavior in deer (Verme & Ullrey, 1984; see below). This latter was reflected by the poor nutritional condition of lactating adult females in the study area (i.e., <5.7% body fat annually at the annual peak in late autumn; Bender et al., 2011, 2013), and in the arid Southwest in general (Bender et al., 2007, 2011, 2012; Bender, 2020), as well as the significant losses in condition seen over winter (Bender & Hoenes, 2017; Hoenes, 2008), despite mule deer likely requiring lower quality forage than white-tailed deer (O. virginianus) (Staudenmaier et al., 2022).
While lack of deep snow cover may result in relative forage availability being less impacted during winter in the arid Southwest, senesced forages are still of very low quality, similar to northern environments (Kemp, 1983; Krausman et al., 1990; McKinney, 2003). Low forage quality in winter was reflected in the condition dynamics of deer on the CRLRC; for example, dry females lost on average 32% of body fat reserves and 38% of rump body condition score over winter (L. Bender, unpublished data), even though their condition in late autumn-early winter was already low relative to other southwestern populations (Bender et al., 2007, 2011, 2012). The nutritional condition of deer is driven primarily by forage quality (Bender, 2020; National Research Council, 2007; Tollefson et al., 2010; Verme & Ullrey, 1984; Wakeling & Bender, 2003), illustrating that deer in the arid Southwest face similar constraints in terms of limited nutrient gains from forage intake as do deer in northern environments during winter.
Table 3
Mean movement distances (x) and mean differences (D) in mean distance moved daily as predicted by continuous-time movement modeling (daily) and subdaily straight line movements between successive 4 h locations (subdaily) of adult female mule deer for which movement data for successive seasons were available between winter (Dec.-Feb.) and spring-autumn (Mar.-Nov.) on the Corona Range and Livestock Research Center, east-central New Mexico, 2005-2007. p = Probability that seasonal differences differ from 0; N = number of seasonal comparisons.
Period | x Summer | x Winter | D | 90% CI | p | N |
Subdaily | 400.2 | 309.6 | -90.6 m | -123- -63 | 1.000 | 20 |
Daily | 4.21 | 3.16 | -1.05 km | -1.53- -0.63 | 1.000 | 20 |
Deer in northern environments do face additional energetic challenges associated with persistent snow cover, which can limit forage availability (by making location and acquisition of food more difficult; Hovey & Harstad, 1992), diet quality (due to reduced forage quality and availability; McKinney, 2003; Osborn & Jenks, 1998), and increase loss of endogenous reserves (because of increased costs of moving through snow; Bunnell et al., 1990; Mattfeld, 1973). However, while deer in the arid Southwest are less influenced by snow-depth related challenges, northern deer do not face the lack of free water experienced by most deer populations in arid environments because of persistent snow cover in northern environments. Lack of water can present an energetic cost to deer in the arid Southwest, as mule deer may increase movement distances to access water (Heffelfinger, 2006), and winter is much drier than spring-autumn in Chihuahuan desert and short-grass prairie habitats of the arid Southwest (e.g., 87% of precipitation occurs during spring-autumn on the CRLRC).
Because of permanent water developments, water was comparably available seasonally on the CRLRC; water developments were accessible from all deer home ranges, so mule deer did not need to alter their movements in response to seasonal changes in availability of water. Hence, need or preference for free water likely had a negligible effect on deer movements on the CRLRC, unless presence of temporary sources (ephemeral pools, etc.) during the summer monsoon reduce deer movements during spring-autumn because of increased availability. Thus, despite water being effectively controlled in our study, deer still showed less movements during winter. This again supports decreased movements during winter being most influenced by the lack of energetic benefit from seeking and foraging on low quality senesced forage.
Cold temperatures are often thought to influence the winter conservation strategy, despite demonstrated lack of benefit of thermal cover in winter to deer (Cook et al., 1998; Freddy, 1984), the pronounced effect of solar radiation on warming deer (Cook et al., 1998; Parker & Gillingham, 1990; Parker & Robbins, 1984), and deer movements and tolerance of exposure to cold (including bedding in the open) except during the most extreme conditions when high quality forage is available (Moen, 1968; Verme & Ullrey, 1984). Mule deer possess a low thermal critical zone (ca. -20 oC) and show greater tolerance of cold than do white-tailed deer, including a lower metabolic rate response to decreasing temperatures (Mautz et al., 1985; Parker & Robbins, 1984). Hence, they are less affected by even extreme cold, which is seldom the case in the arid Southwest where mean low temperatures seldom approach their lower thermal tolerance (Table 1). Conversely, the upper thermal critical level of mule deer in winter pelage (5 ºC; Mautz et al., 1985) is lower than average high temperatures during winter on our study area (6.7-8.9 ºC; Table 1) and much of the arid Southwest. This would require active metabolic activity or behaviors (e.g., panting, etc.) by mule deer to cool themselves, or possibly limiting movements and bedding under shade (although the energetic benefits of the latter are questionable; Cook et al., 1998). Consequently, if temperature affected movements during the winter in our study area, deer would be more likely to decrease movements because of heat stress, not cold stress.
Lastly, while at least one behavioral aspect of the winter energy conservation strategy is seen in mule deer in the arid Southwest, to what degree other adaptations are conserved is unknown. Mule deer in the arid Southwest do develop a highly insulative pelage in winter (Heffelfinger, 2006), but the extent that they may decrease metabolic rate or regulate vascular circulation to the extremities (or need to, in light of the more moderate temperatures) is unknown. Additionally, while lactating females showed a tendency to move less relative to dry females on the CRLRC, whether this is typical is unknown because of our small sample of lactating females. However, given that most females on CRLRC were in poorer condition than females elsewhere in the arid Southwest (Bender et al., 2007, 2011, 2012; Bender & Hoenes, 2017), if very low condition (such as results from lactation) increases movements during winter, this increase would likely have been seen in all CRLRC females regardless of lactation status. Moreover, GPS/VHF and VHF-only collared females showed similar movements (as indexed by home range sizes) seasonally and annually on CRLRC, indicating that movements of GPS/VHF collared females reflected females in general. Although small sample sizes (3-4) precluded including lactation status as an interactive term in the contrast of GPS/VHF and VHF-only collared females, for all females lactation status had no effect on annual or seasonal home range sizes (p ≥ 0.599).
Acknowledgments
Support for this project was provided by the U.S. Forest Service-Rocky Mountain Research Station and the New Mexico State University Cooperative Extension Service and Agricultural Experimental Station. All activities were in accordance with NMSU IACUC Permit No. 2005-023. E.J.T.-R. was supported by a postdoctoral fellowship from Consejo Nacional de Humanidades, Ciencias y Tecnologías (Conahcyt-Mexico).
References
Alldredge, A. W., Lipscomb, J.F., & Whicker, F.W. (1974). Forage intake rates of mule deer estimated with fallout
cesium-137. The Journal of Wildlife Management, 38, 508–516. https://doi.org/10.2307/3800882
Bartmann, R. M., & Bowden, D. C. (1984). Predicting mule deer mortality from weather data in Colorado. Wildlife Society Bulletin, 12, 246–248. https://doi.org/10.1002/jwmg.956
Bender, L. C. (2020). Elk, deer, and pinyon-juniper: needs, what works, and what doesn’t. In K. Malcolm, B. Dykstra, K. Johnson, D. Lightfoot, E. Muldavin, & M. Ramsey (Compilers), Symposium proceedings on piñon-juniper habitats: status and management for wildlife – 2016. Proceedings RMRS-P-77. Fort Collins, Colorado: U.S. Forest Service, Rocky Mountain Research Station.
Bender, L. C., & Hoenes, B. D. (2017). Costs of lactation to body condition and future reproduction of free-ranging mule deer Odocoileus hemionus (Cervidae). Mammalia, 81, 329–338. https://doi.org/10.1515/mammalia-2015-0143
Bender, L. C., Lomas, L. A., & Browning, J. (2007). Condition, survival, and cause-specific mortality of mule deer in northcentral New Mexico. Journal of Wildlife Management, 71, 1118–1124. https://doi.org/10.2193/2006-226
Bender, L. C., Boren, J. C., Halbritter, H., & Cox, S. (2011). Condition, survival, and productivity of mule deer in semiarid grassland-woodland in east-central New Mexico. Human-Wildlife Interactions, 5, 276–286. https://doi.org/10.26077/60n0-ks48
Bender, L. C., Hoenes, B. D., & Rodden, C. L. (2012). Factors influencing survival of desert mule deer in the greater San Andres Mountains, New Mexico. Human-Wildlife Interactions, 6, 245–260. https://doi.org/10.26077/h5bg-1829
Bender, L. C., Boren, J. C., Halbritter, H., & Cox, S. (2013). Effects of site characteristics, pinyon-juniper control, and precipitation on habitat quality for mule deer on the Corona Range and Livestock Research Center. Human-Wildlife Interactions, 7, 47–59. https://digitalcommons.usu.edu/hwi/vol7/iss1/5
Bunnell, F. L., Hovey, F. W., McNay, R. S., & Parker, K. L. (1990). Forest cover, snow conditions, and black-tailed deer sinking depths. Canadian Journal of Zoology, 68, 2403–2408. https://doi.org/10.1139/z90-333
Calabrese, J. M., Fleming, C. H., & Gurarie, E. (2016). ctmm: an R package for analyzing animal relocation data as a continuous-time stochastic process. Methods in Ecology and Evolution, 7, 1124–1132. https://doi.org/10.1111/2041-210X.12559
Calabrese, J. M., Fleming, C. H., Noonan, M. J., & Dong, X. (2021). ctmmweb: a graphical user interface for autocorrelation-informed home range estimation. Wildlife Society Bulletin, 45, 162–169. https://doi.org/10.1002/wsb.1154
Cook, J. G., Irwin, L. L., Bryant, L. D., Riggs, R. A., & Thomas, J. W. (1998). Relations of forest cover and condition of elk: a test of the thermal cover hypothesis in summer and winter. Wildlife Monographs, 141, 1–61.
Efron, B., & Tibshirani, R. J. (1993). An introduction to the bootstrap. New York: Chapman & Hall.
Fleming, C. H., Fagan, W. F., Mueller, T., Olson, K. A., Leimgruber, P., & Calabrese, J. M. (2014). From fine-scale foraging to home ranges: a semivariance approach to identifying movement modes across spatiotemporal scales. American Naturalist, 183, E154–E167. http://doi.org/10.1086/675504
Fleming, C. H., Fagan, W. F., Mueller, T., Olson, K. A., Leimgruber, P., & Calabrese, J. M. (2016). Estimating where and how animals travel: an optimal framework for path reconstruction from autocorrelated tracking data. Ecology, 97, 576–582. https://doi.org/10.1890/15-1607.1
Freddy, D. J. (1984). Quantifying capacity of winter ranges to support deer – evaluation of thermal cover used by deer. Denver, CO: Wildlife Research Report, Colorado Division of Wildlife, USA.
Heffelfinger, J. (2006). Deer of the Southwest: a complete guide to the natural history, biology, and management of Southwestern mule deer and white-tailed deer. Texas A&M University, College Station, Texas.
Hoenes, B. D. (2008). Identification of factors limiting desert mule deer populations in the greater San Andres Mountains of southcentral New Mexico (Thesis). New Mexico State University, Las Cruces.
Hovey, F. W., & Harestad, A. S. (1992). Estimating effects of snow on shrub availability for black-tailed deer in southwestern British Columbia. Wildlife Society Bulletin, 20, 308–313.
Jacobsen, N. K. (1980). Differences of thermal properties of white-tailed deer pelage between seasons and body regions. Journal of Thermal Biology, 5, 151–158. https://doi.org/10.1016/0306-4565(80)90014-5
Kemp, P. R. (1983). Phenological patterns of Chihuahuan desert plants in relation to the timing of water availability. Journal of Ecology, 71, 427–436.
Kie, J. G., Baldwin, J. A., & Evans, C. J. (1996). CALHOME: a program for estimating animal home ranges. Wildlife Society Bulletin, 24, 342–344.
Krausman, P. R., Ordway, L. L., Whiting, F. M., & Brown, W. H. (1990). Nutritional composition of desert mule deer forage in the Picacho Mountains, Arizona. Desert Plants, 10, 32–34.
Marchinton, R. L., & Hirth, D. H. (1984). Behavior. In L. K. Halls (Ed.), White-tailed deer ecology and management. Pennsylvania: Stackpole Books.
Marshal, J. P., Krausman, P. R., & Bleich, V. C. (2008). Body condition of mule deer in the Sonoran Desert is related to rainfall. Southwestern Naturalist, 53, 311–318. https://doi.org/10.1894/CJ-143.1
Mattfeld, G. F. (1973). The effect of snow on the energy expenditure of walking white-tailed deer. In Transactions of the 30th Northeast Fish and Wildlife Conference, Dover, Vermont, Spring, April 1973.
Mautz, W. W., Pekins, P. J., & Warren, J. A. (1985). Cold temperature effects on metabolic rates of white-tailed, mule, and black-tailed deer in winter coat. In P. F. Fennessy, & K. R. Orew (Eds.), The biology of deer. Royal Society of New Zealand Bulletin, 22, 453–457.
McKinney, T. (2003). Precipitation, weather, and mule deer. In J. C. Jr. deVos, M. R. Conover, & N. E. Headrick (Eds.), Mule deer conservation. Issues and management strategies. Logan, Utah: Jack H. Berryman Press.
Moen, A. N. (1968). Energy exchange of white-tailed deer, western Minnesota. Ecology, 49, 676–682. https://doi.org/10.2307/1935531
National Research Council (US). (2007). Committee on Nutrient Requirements of Small Ruminants. Nutrient requirements of small ruminants: sheep, goats, cervids, and New World camelids. Washington D.C.: National Academy Press.
Osborn, R. G., & Jenks, J. A. (1998). Assessing dietary quality of white-tailed deer using fecal indices: effects of supplemental feeding and area. Journal of Mammalogy, 79, 437–447. https://doi.org/10.2307/1382974
Parker, K. L., & Robbins, C. T. (1984). Thermoregulation in mule deer and elk. Canadian Journal of Zoology, 62, 1409–1422. https://doi.org/10.1139/z84-202
Parker, K. L., & Gillingham, M. P. (1990). Estimates of critical thermal environments for mule deer. Journal of Range Management, 43, 73–81. http://doi.org/10.2307/3899126
Robinette, W. L., Jones, D. A., Rogers, G., & Gashwiler, J. S. (1957). Notes on tooth development and wear for Rocky Mountain mule deer. Journal of Wildlife Management, 21, 134–153. https://doi.org/10.2307/3797579
SAS. (1988). SAS/STAT user’s guide. SAS Institute Incorporated, Cary, North Carolina.
Short, H. L. (1981). Nutrition and metabolism. In O. C. Wallmo, R. E. McCabe, & L. R. Jahn (Eds.), Mule and black-tailed deer of North America. Lincoln: University of Nebraska Press.
Silver, H., Holter, J. B., Colovox, N. F., & Hayes, H. H. (1971). Effect of falling temperature on heat production in fasting white-tailed deer. Journal of Wildlife Management, 35, 37–46.
Staudenmaier, A. R., Shipley, L. A., Camp, M. J., Forbey, J. S., Hagerman, A. E., Brandt, A. E. et al. (2022). Mule deer do more with less: comparing their nutritional requirements and tolerances with white-tailed deer. Journal of Mammalogy, 103, 178–195. https://doi.org/10.1093/jmammal/gyab116
Tollefson, T. N., Shipley, L. A., Myers, W. L., Keisler, D. H., & Dasgupta, N. (2010). Influence of summer-autumn nutrition on body condition and reproduction in lactating mule deer. Journal of Wildlife Management, 74, 974–986. https://doi.org/10.2193/2008-529
Verme, L. J., & Ullrey, D. E. (1984). Physiology and nutrition. In L. K. Halls (Ed.), White-tailed deer: Ecology and management. Harrisburg, Pennsylvania: Stackpole Books.
Wakeling, B. F., & Bender, L. C. (2003). Influence of nutrition on mule deer biology and ecology. In J. C. Jr. deVos, M. R. Conover, & N. E. Headrick (Eds.), Mule deer conservation. Issues and management strategies. Logan, Utah: Jack H. Berryman Press.
Impacts of disturbance on ant (Hymenoptera:Formicidae) food preferences and dominance in a Mexican temperate forest.
Meghan I. Zolá-Rodríguez a, Mariana Cuautle b, *, Marco Daniel Rodríguez-Flores c, Citlalli Castillo-Guevara b
a Akumal Monkey Sanctuary & Rescued Animals, Camino a Uxuxubi s/n Predio Santa Pilar Lote 16, 77776 Akumal, Quintana Roo, Mexico
b Universidad Autónoma de Tlaxcala, Centro de Investigación en Ciencias Biológicas, Km 10.5 Carretera Tlaxcala-San Martín Texmelucan, 90120 San Felipe Ixtacuixtla, Tlaxcala, Mexico
c Blue Marlin Conservation (Conservation Diver), Sunset Beach Gili Air, Gili Indah, Pemenang, North Lombok Regency, West Nusa Tenggara 83355, Indonesia
*Corresponding author: mcuautle2004@gmail.com (M. Cuautle)
Received: 26 June 2024; accepted: 27 August 2024
Abstract
This study examines the impact of disturbance on the food preferences and dominance of an ant community in a temperate ecosystem in Mexico. The study focused on 2 types of vegetation: native oak forest and induced grassland (disturbed vegetation). Observations were conducted to record the food elements carried by ants to their nests. These data were analyzed using x2 tests. Tuna and honey baits were placed near the nests to record the presence of ants in 5-minute periods. We used a binomial model to determine whether the probability of finding an ant foraging at the baits was affected by vegetation type, bait type, and/or ant species. Additional baits were used to determine the ant dominance indices. T-tests and ANOVAs were used to compare dominance indices between vegetation types, baits, and ant species. No significant differences were observed in food preferences between vegetations. However, some species showed a preference for honey (i.e., carbohydrates), which could be limited in ground-level environments. Ants showed a submissive behavior in both vegetation types. This research shows that ants could optimize their nutrient intake, enabling them to survive efficiently even when facing disturbances, instead of increasing dominance.
Keywords: Ant nest; Dominance index; Feeding habits; Compensation hypothesis; Carbohydrates; Proteins
© 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/).
Impacto del disturbio en las preferencias alimentarias y dominancia de las hormigas (Hymenoptera: Formicidae) en un bosque templado de México
Resumen
Este estudio examina el impacto del disturbio en las preferencias alimentarias y dominancia de una comunidad de hormigas en un ecosistema templado en México, en bosque de encino nativo y pastizal inducido (vegetación perturbada). Se registraron los alimentos transportados por las hormigas a sus nidos. Estos datos fueron analizados utilizando pruebas de x2. Se colocaron cebos de atún y miel cerca de los nidos para registrar la presencia de hormigas. Utilizamos un modelo binomial para determinar si la probabilidad de encontrar una hormiga en los cebos se veía afectada por la vegetación, cebo o especie de hormiga. Los índices de dominancia se determinaron usando cebos. Se emplearon pruebas t y Anova para comparar los índices de dominancia entre tipos de vegetación, cebos y especies de hormigas. No hubo diferencias significativas en las preferencias alimentarias entre tipos de vegetación, pero algunas especies mostraron una preferencia por la miel (carbohidratos), que podría ser un recurso limitado a nivel del suelo. Las hormigas mostraron un comportamiento sumiso en ambos tipos de vegetación. Esta investigación muestra que las hormigas podrían optimizar su ingesta de nutrientes, permitiéndoles sobrevivir bajo condiciones de disturbio, en lugar de aumentar su dominancia.
Palabras clave:Nidos de hormigas; Índice de dominancia; Hábitos alimenticios; Hipótesis de compensación; Carbohidratos; Proteínas
Introduction
Land use change stands as one of the primary factors contributing to global change and its adverse effects on biodiversity (Ellis et al., 2010; Foley et al., 2005; Sala et al., 2000). Land use change has reduced biodiversity through the loss, modification, and fragmentation of habitats; degradation of soil and water; and overexploitation of native species (Foley et al., 2005), and its effects depend strongly on the type, severity, frequency and timing of disturbance (Foley et al., 2005; White & Jentsch, 2001). Ants are one of the most dominant insects both ecologically and numerically (Rico-Gray & Oliveira, 2007; Schultheiss et al., 2022; Toro et al., 2012) that participate in different ecological processes —e.g., nutrient recycling, soil formation, decomposition, seed dispersion (Toro et al., 2012). Ants are model organisms to study the effect of disturbance because they respond to environmental change (Agosti et al., 2000; Andersen, 2000). It is often observed that disturbances favor behaviorally dominant ants, including invasive species. Vonshak and Gordon (2015), observed that native ant richness was highest in natural habitats, and alien species richness was highest in urban habitats, along an urban-rural gradient in the San Francisco Bay area. Nevertheless, the specific effect will depend on factors like the type of disturbance, including fires, floodings, or even treefall gaps (Cerdá et al., 2013). Hoffman and Andersen (2003) studied the response of ant functional groups to disturbance finding that the dominant Dolichoderinae and “hot climate specialists” tend to be favoured by low levels of disturbance. The opportunist and generalized Myrmicinae have wide habitat tolerances but are sensitive to competitive interactions, while “cryptic species” and “specialist predators” are highly sensitive to disturbance.
Feeding habits and foraging strategies represent vital life history traits among ants, playing a significant role in their ecological dynamics. These traits are contingent upon the availability of resources within their respective habitats (Andersen, 2000; Davidson, 2005), and the ensuing competition for those resources (Arnan et al., 2012). The feeding habits of ants play a pivotal role in determining their ecological function within the ecosystem (Spotti et al., 2015). However, despite their significance, the food habits and preferences of ants remain relatively understudied and represent one of the least understood aspects of their biology (Houdria et al., 2015).
The effects of disturbance (e.g., land use change) on ant feeding habits (especially foraging strategies) have not been studied in depth (Castillo-Guevara et al., 2019; Hernández-Flores et al., 2016; Radnan et al., 2018), and the way ants respond to habitat complexity can provide clues as to how ants could respond to disturbance. When the external disturbing factors are of high intensity (natural-fires or anthropogenic-deforestation), the disturbances can initiate a directional regression manifested as gradual or rapid simplification of the horizontal and vertical structure of a community, leading to the replacement of complex communities by a simpler one (Łaska, 2001). The level of habitat complexity can influence ant foraging patterns. For instance, higher complexity may decrease interspecific interactions and ant recruitment or minimize the trade-off between resource discovery and dominance (Parr & Gibb, 2012).
Examining the effect of soil surface complexity on food exploitation in the context of change from grassland to shrubland in Australia, Radnan et al. (2018) discovered that substrate complexity (wood debris, leaf litter, or no substrate) influenced the discovery time, ant size, and monopolization index of tuna and honey baits within testing arenas. However, at a larger scale of vegetation type, the effect was not observed. Similarly, Castillo-Guevara et al. (2019) found that the dominance level was similar in natural oak vegetation in comparison to an agricultural area in a temperate community; however, ant foraging strategies differed between the 2 communities.
Competition has been regarded as the primary factor influencing the structure of ant communities (Cerdá et al., 2013). Through the use of aggressive strategies, ants exert influence on the spatial distribution, abundance, and behavior of other ants. These strategies encompass a range of tactics, including the deployment of repellent chemicals and the establishment of territories (Cerdá et al., 2013). Moreover, ants can establish interspecific dominance hierarchies primarily based on variations in food collection behavior and aggressiveness. While there are various proposals for categorizing such hierarchies, the classification proposed by Vepsäläinen and Pisarski (1982) and Savolainen et al. (1989), as described by Cerdá et al. (2013), is considered the most well-defined hierarchical system from an ecological perspective. This classification system is founded on the aggressive behavior of ants and its impact on other ant species, and includes 3 categories: the dominant, subdominant and subordinate category. Dominant species exhibit highly aggressive behavior, exhibit numerical predominance over other species, fiercely defend their territories, and establish mutual exclusivity within their ecological communities. Subdominant species, while not actively defending territories, display a remarkable propensity for aggressively safeguarding their food resources (Cerdá et al., 2013). Lastly, subordinate species, characterized by small colonies devoid of recruitment systems, employ strategies to avoid physical confrontations with other colonies and species. Nonetheless, they exhibit a strong commitment to defending their nests against potential threats (Cerdá et al., 2013). Different foraging strategies, such as the subordinate species’ ability to discover resources before dominant species (dominance-discovery trade-off, Fellers, 1987) or their capacity to forage across a broader temperature range compared to dominant species —dominance-thermal tolerance trade-off (Fellers, 1989), can contribute to the coexistence of ants within a community.
The present study focuses on the analysis of ant feeding habits and dominance hierarchy in response to land use change within a temperate ecosystem located in central Mexico. A habitat with the native oak forest vegetation (complex habitat), was compared to a nearby area where the land use had been altered to induced grassland (simplified habitat). The study hypothesized: i) it is expected that in the oak forest, food items transported to the nest will be more varied than in the induced grassland. This variation at the species level is also anticipated due to the change in land use. The conversion from a complex habitat (oak forest) to a simplified habitat (induced grassland) will likely provide fewer resources variety for the ants; ii) it is expected that as a result of the disturbance in the induced grassland, ants will be more generalized in low heterogeneous habitats due to the dominance of generalist ant species. Additionally, the arrival times at the baits will be shorter due to the absence of leaf litter; iii) a lower dominance index is expected in the induced grassland than in the oak forest, due to the presence of subdominant species and the modification of the dominance hierarchy system by the land use change.
Materials and methods
The study was conducted within “Flor del Bosque” State Park, located in the municipality of Amozoc de Mota, in the state of Puebla. The coordinates of this protected reserve are 19°00’00”-19°01’50” N, 98°20’35”-98°20’53” W. The State Park encompasses an area of 664.03 hectares, characterized by altitudes ranging from 2,225 to 2,400 m asl. The annual average temperature in this region fluctuates between 14 °C and 16 °C, with the majority of rainfall occurring during the summer months, contributing to an average annual precipitation of 750 to 950 mm. It is important to note that the park experiences a distinct dry season lasting approximately 6 months, from November to April, as reported by Costes et al. (2006). The native vegetation of the reserve primarily consists of oak forest [Quercus castanea (Née), Q. laurina (Bonpl.), Q. laeta (Liebm.) (Fagaceae)]. However, human disturbances have led to the presence of induced grassland areas and, to a lesser extent, eucalyptus plantations [Eucalyptus spp. L´Hér) (Myrtaceae)] within the park, as documented by Costes et al. (2006). Given that the disturbed vegetation within the reserve primarily consisted of induced grassland, our research aimed to establish a meaningful comparison between native and altered vegetation.
Ant communities were surveyed once a month during specific periods (April, August and October 2015, and January to March 2016). The survey methodology involved the establishment of 6 transect plots (400 m × 20 m), with 3 plots located in the oak forest and 3 in the induced grassland. The spatial distribution of these transects can be observed in Figure 1. To locate ant nests, the transects were traversed, and various substrates such as leaf litter, stones, trunks, and branches were lifted and examined. Additionally, a total of 5 tuna baits and 5 honey baits were placed along the transects at 10 m intervals to attract ants and to follow them to locate their nest. The nests found during the survey were georeferenced for accurate spatial documentation. A 5-minute observation period was designated for each nest, wherein all food items transported by the ants to their nests were recorded. These observations were conducted between 9:00 a.m. and 3:00 p.m. on the sampling days. Each recorded item was subsequently categorized into one of the following groups: 1) plant elements, encompassing any component of plants except for seeds, 2) seeds, and 3) arthropods. After each observation period, 2 bait samples, enclosed within Petri dishes, were positioned in close proximity to the nests (approximately 10 cm), maintaining a distance of approximately 5 centimeters from each other. Various studies exploring the feeding habits, preferences, and foraging strategies of ants have employed diverse bait types. Typically, these baits consist of carbohydrates such as honey and other sources representing proteins such as tuna (Houdria et al., 2015; Lynch et al., 1980; Spotti et al., 2015; Trigos-Peral et al., 2016). The bait materials employed in the study consisted of honey and tuna, which were consistently provided in a standardized amount of one tablespoon, equivalent to approximately 5 grams. To quantify the number of individual ants on each type of food resource, we placed baits and observed them for 5 min. Ants responded very quickly to the baits; therefore, the observation time was limited to 5 min. Within this duration, the number of individuals was recorded for each bait type. After the observation periods, 1 to 3 ants were collected from each nest using a vacuum cleaner or tweezers and then preserved in Eppendorf tubes containing 70% ethanol. In the laboratory, the ants were separated, mounted, and identified to the genus level using the key by Mackay and Mackay (1989). In certain instances, the species identification was feasible by comparing the collected specimens with those present in the Entomological Collection of the Universidad de las Américas Puebla (UDLAP).
To analyze the food preferences of the ant community, we considered 3 factors. First, we determined whether the number of foragers carrying plants, seeds, or arthropods to the nest differed between the whole ant communities of each vegetation type or among ant species. We used the appropriate contingency tables and chi-square tests for such comparisons. Second, we determined whether the presence of foragers at each bait during each minute of the 5-minute period was influenced by vegetation type, bait type (honey or tuna), ant species or interactions among these factors. We used a binomial model to evaluate this response variable, with “ant presence” coded as 1 and “absence” as 0. To select the most relevant explanatory variables, we employed a stepwise forward approach. We began with the simplest model and successively added each response variable. The significance of adding each variable was assessed using an x2 test, comparing the previous model with the new model. If a variable was found to be statistically significant, it was retained in the model. Otherwise, it was removed, and we proceeded with the next explanatory variable. This stepwise procedure allowed us to build the final model with the most significant factors explaining the presence of foragers at each bait. No overdispersion was detected in the selected model. Post-hoc analyses were conducted when significant differences were detected. The statistical analysis of data was performed using R software (R Core Team, 2022).
We evaluated the dominance hierarchy of the ants in each vegetation type (oak forest and induced grassland), ant species, and bait types using the same transects as for the detection of the nests (Fig. 1), during October 2015 and January, February and March (2016). At each transect, 9 sampling points were established on the ground, spaced 10 m apart. Each sampling point consisted of a pair of Petri dishes with baits: one with honey and one with tuna, placed less than 5 cm apart. This resulted in a total of 540 baits (5 replicates × 6 transects × 9 sampling points × 2 baits; 2 samplings in February). These types of baits have been widely used for ant dominance hierarchy studies (Dáttilo et al., 2014; Parr & Gibb, 2012; Trigos-Peral et al., 2016).
For the first 3 sets of bait, we recorded the arrival times of the ants during a one-hour observation period. Most of the ant species were identified in the field. For the ants that were not identified, a few individuals (2 to 3 specimens) were collected. In the case of the remaining 6 sets of bait, they were filled with water. Ants that fell onto the Petri dish after 2 h were collected. The presence of water did not deter the ants from visiting, but it allowed us to determine which ants had been attracted to the baits. The baits were placed and retrieved from different transects in the field between 9:00 a.m. and 5:00 p.m. The order of the transects was changed on different sampling days to mitigate the potential impact of the time of day. Ant specimens collected from the baits were preserved in 70% ethanol and transported to the laboratory for further identification.
Figure 1. Localization of the transects used in the study zone (mapped by Luna F.).
We chose numerical dominance to assess the dominance hierarchy of the ants —i.e., ordering of ant species based on their numerical or behavioral dominance by vegetation types (Andersen, 1992; Cerdá et al., 1997; Stuble et al., 2017), ant species and bait type. As numerical and behavioral dominance are highly correlated, this method has been well-established and documented in the ant literature (Dáttilo et al., 2014; Dejean & Corbara, 2003; Parr, 2008; Parr & Gibb, 2012; Santini et al., 2007). This method indicates which species are consistently present at the baits, and which ones dominate the baits numerically and thus monopolize them (Parr, 2008). We represented numerical dominance using the numerical dominance index (DI) for each morphospecies calculated by the formula: DI = (Di)/(Di + Si), where, Di is the number of baits monopolized by the species of ant i, and Si is the number of baits that the species of ant i used but did not monopolize. Baits were considered to be monopolized when more than 5 individuals (workers and/or soldiers) of the same morphospecies were using the resource without the presence of other morphospecies. This measure (more than 5 individuals) takes into consideration that in temperate climates, ants are less abundant, and recruitment is considered weaker than in tropical environments where the index has been widely more (Santini et al., 2007). Therefore, dominant morphospecies are those that find and monopolize a larger proportion of the food resources in a given environment. The value of the index ranges from 0 (completely submissive species) to 1 (totally dominant species) and is similar to the “monopolization index” used in other studies (Dáttilo et al., 2014; Fellers, 1987; Parr & Gibb, 2012; Santini, et al., 2007). In this study, ant species with a DI lower than 0.5 were classified as submissive.
The arrival times of the ants were compared using survival curves using the survival package of the R Software program. The one-hour and 2-hour baits were used to calculate the ID of the ant species. To compare the ID between vegetation and bait types, t-tests were performed. To compare the DI between ant species, a one-way ANOVA test was performed after applying a square root transformation to meet the normality requirements. These analyses were performed using the program STATVIEW 5.0 (Abacus Concepts Inc., 1996).
Results
Fourteen morphospecies of ants were recorded in this study belonging to 11 ant genera (Table 1). For the analysis of the results from the nests and the baits placed near them, 2 morphospecies of Pheidole were identified. However, due to the low number of records for each morphospecies across different vegetation types, Pheidole sp. 1 and Pheidole sp. 2 were combined for the analysis. In the oak forest, 2 out of 19 records corresponded to Pheidole sp. 1, and in the grassland, 4 out of 15 records corresponded to Pheidole sp. 2. Since Pheidole species often share similar feeding habits and are functionally similar (Agosti, 2000; Andersen, 2000), these 2 morphospecies were grouped to increase statistical power, and the analysis was conducted at the genus level. In contrast, another Pheidole species, Pheidole sp. 3, was recorded in the baits used to determine the DI and was analyzed independently due to its different context of occurrence. The rest of the analyses were conducted at the species level.
In the case of the nests, we found 9 morphospecies distributed among 34 nests, 19 of these were located in the oak forest and 18 in the induced grassland (Table 2). No significant differences were observed in the types of food elements those individual ants transported to their nest when comparing different vegetation types (x2 = 2.13, df = 2, p = 0.34). In the oak forest considering all ant species, there were no significant differences in the number of individuals carrying elements from different categories (x2 = 0.03, df = 1, p = 0.86) (Fig. 2). Of the genera of ants observed carrying food elements to their nests, only 2 genera were observed with a single type of food resource: Prenolepis imparis (Say) only carried arthropods and Dorymyrmex insanus (Buckley) only seeds.
Table 1
List of ant species recorded in this study in “Flor del Bosque” State Park in Puebla, Mexico.
Formicidae/Subfamily/Ant species | |||
Formicidae Latreille, 1809 | |||
Dolichoderinae Forel, 1878 | |||
Leptomyrmecini Emery, 1913 | |||
Dorymyrmex Mayr, 1866 | |||
Dorymyrmex insanus (Buckley, 1866) | |||
Linepithema Mayr, 1866 | |||
Linepithema dispertitum (Forel, 1885) | |||
Dorylinae Leach, 1815 | |||
Dorylini Ashmead, 1905 | |||
Labidus Jurine, 1807 | |||
Labidus coecus Latreille, 1802 | |||
Formicinae Latreille, 1809 | |||
Camponotini Forel, 1878 | |||
Camponotus Mayr,1861 | |||
Camponotus rubrithorax Forel, 1899 | |||
Lasiini Ashmead, 1905 | |||
Nylanderia Emery, 1906 | |||
Nylanderia austroccidua Trager, 1984 | |||
Prenolepis Mayr, 1861 | |||
Prenolepis imparis Say, 1836 | |||
Myrmicinae Lepeletier, 1835 | |||
Attini Smith, 1858 | |||
Pheidole Westwood, 1839 | |||
Pheidole sp. 1 | |||
Pheidole sp. 2 | |||
Pheidole sp. 3 | |||
Crematogastrini Forel, 1893 | |||
Temnothorax Mayr, 1861 | |||
Temnothorax augusti Baroni Urbani, 1978 | |||
Pogonomyrmecini Ward, Brady, Fisher & Schultz, 2014 | |||
Pogonomyrmex Mayr, 1868 | |||
Pogonomyrmex barbatus (Smith, F., 1858) | |||
Solenopsidini Forel, 1893 | |||
Monomorium Mayr, 1855 | |||
Monomorium ebenium Forel, 1891 | |||
Pseudomyrmecinae, Smith, M.R., 1952 | |||
Pseudomyrmecini Smith, 1952 | |||
Pseudomyrmex Lund, 1831 | |||
Pseudomyrmex pallidus Smith, F., 1855 |
Conversely, ants of the genus Pheidole spp. carried a greater quantity of plant elements compared to the other categories (x2 = 10.4, df = 2, p = 0.006); only 2 of these individuals were observed carrying arthropods, and none were recorded carrying seeds. No significant differences were found for D. insanus (x2 = 2, df = 2, p = 0.37), P. barbatus (x2 = 1.6, df = 2, p = 0.450), nor P. imparis (x2 = 4, df = 2, p = 0.135) among the 3 food categories. Only P. barbatus carried all 3 types of food (Fig. 2).
Considering all ant species, there were no significant differences in the number of individuals carrying different food elements in the induced grassland (x2 = 0.66, df = 1, p = 0.41, Fig. 2). Among the observed ant species that transported food types to their nests, P. barbatus was found to carry both plant elements and seeds. Additionally, Dorymyrmex insanus and Pheidole spp. were observed carrying food from all 3 categories (Fig. 2). Pheidole spp. (x2 = 10.33, df = 2, p = 0.006) and P. barbatus (x2 < 30.33, df = 2, p < 0.001), carried a higher number of plant elements. For D. insanus no significant differences were found in the number of elements of each type (x2 < 3.77, df = 2, p = 0.15) (Fig. 2).
Figure 2. Number of individuals of the different ant species recorded carrying any food resource to their nest. DORY: Dorymyrmex insanus, PHEI: Pheidole spp., POGO: Pogonomyrmex barbatus, PREN: Prenolepis impairs.
In the induced grassland, 5 ant species were attracted to both types of baits, and 1 species was found in the tuna baits. In the oak area, 5 ant species were observed in both types of baits, and 1 species was attracted to the honey bait (Fig. 3). The probability of presence of a forager during a given minute of the observation period depended on the vegetation type (x2= 8.73, df = 1, p = 0.03), the ant species (x2= 59.64, df = 7, p < 0.001), the interaction between the ant species and the bait (x2= 57.96, df = 8, p < 0.001), and the interaction between the vegetation type, the ant species, and the bait (x2= 26.12, df = 7, p < 0.001). The Tukey contrast test for the vegetation*species*bait interaction showed (Z > 3.83, p < 0.042) that the presence of Linepithema dispertitum (Forel) on the honey bait, in the oak forest, was more likely than the presence of Pheidole spp. on either the tuna or honey bait in the oak forest, Pheidole spp. on the honey bait in the induced grassland, and P. impairs on the honey bait in the oak forest (Fig. 3). The probability of Camponothus rubrithorax (Forel) being present on the honey bait in the induced grassland, was greater than that of D. insanus in either the tuna or honey bait in the induced grassland, Pheidole spp. on either the tuna or honey bait in the oak forest, as well as its presence in the honey bait in the induced grassland and P. imparis in the honey bait in the oak forest (Fig. 3). The probability that P. barbatus was present on the honey bait in the induced grassland, was greater than the presence of Pheidole spp. in either the tuna or honey bait and P. imparis in the honey bait in the oak forest (Fig. 3). Finally, the presence of Pheidole spp. on the tuna bait was more likely than its presence on the honey bait in the induced grassland (Fig. 3).
Figure 3. Probability of finding an ant forager on the honey or tuna baits at the nest of the different ant species registered in the Oak Forest (OF) and the Induced Grassland (IG) in the 5-minute period. CAMPO: Camponotus rubrithorax, DORY: Dorymyrmex insanus, LABI: Labidus coecus, LINE: Linepithema dispertitum, MONO: Monomorium ebenium, PHEI: Pheidole spp., POGO: Pogonomyrmex barbatus, PREN: Prenolepis imparis.
Figure 4. Survival curves depicting the arrival times of ant foragers on the baits (tuna or honey) during a one-hour observation.
No significant differences were found in the arrival times among ant species.
No significant differences were found in the arrival times between the vegetation types (x2 = 0.4, df = 1, p = 0.5). Moreover, no significant differences were found in the arrival times among the different ant species (x2 = 7.6, df = 7, p = 0.4) or the type of bait (x2 = 0.01, df = 1, p = 0.9; Fig. 4).
In the dominance index baits 12 ant morphospecies were recorded and different ant morphospecies were recorded in each of the vegetation types (Table 3). When comparing the average dominance index (DI) by vegetation type (mean ± SE, n; oak forest DI = 0.372 ± 0.071, 33; induced grassland DI = 0.251 ± 0.052, 44), was not statistically different (t= 1.091; df = 75; p = 0.2787), which would suggest a submissive behavior in both types of vegetation. Average dominance index by ant species indicates that M. ebenium, P. imparis and Pheidole sp. 3 behave as submissive in both vegetation types (t = 0.590, df = 20, p = 0.5617; t = -0.237, df = 10, p = 0.8171; t = -0.315, df = 11, p = 0.7590, respectively) (Table 4). It was not possible to compare the DI between vegetation types for Temnothorax sp., T. augusti, C. rubrithorax, N. austroccidua, L. dispertitum, D. insanus or P. pallidus, because they were only present in one of the vegetation types (Table 4).
The average dominance index for each ant species was low suggesting that all of them displayed submissive behaviors (F8,68 = 1.949, p = 0.06). Nylanderia austroccidua and L. dispertitum had the highest dominance indices, although it should be noted that N. austroccidua only had one record and L. dispertitum had 3 records. Additionally, Monomorium ebenium showed a tendency to behave as dominant (Table 4).
When comparing the mean dominance index by bait type, no significant differences were found (t = 1.023, df = 110, p = 0.3088; mean ± SE, n; tuna = 0.382 ± 0.060, 55; honey = 0.295 ± 0.053, 57). This indicates that ants exhibited submissive behavior in both types of bait.
According to the average dominance index per ant species in relation to the bait type, Pheidole sp. 3, P. imparis, M. ebeninum, C. rubrithorax, and D. insanus exhibited a submissive behavior in both types of baits (t = 0.888, df = 13, p = 0.3907; t= 0.988, df = 15, p = 0.3388; t = 0.189, df = 32, p = 0.8513; t = -0.918, df = 24, p = 0.3675, t = 2.390, df = 5, p = 0.0624, respectively) (Table 5) and L. dispertitum exhibited a dominant behavior in both types of baits (t = 0.421, df = 3, p = 0.7021).
Temnothorax sp. and P. pallidus were not compared statistically due to the low incidence registered; nonetheless, they presented a submissive behavior index in both types of bait. Neither T. augusti nor N. austroccidua were included in this analysis, as they were only present on the honey bait. In none of the species was there a difference in behavior between the baits.
Table 2
Number of nests found in the oak forest and the induced grassland.
Species | Oak forest | Induced grassland |
Camponotus rubrithorax | 0 | 3 |
Dorymyrmex insanus | 4 | 6 |
Labidus coecus | 1 | 0 |
Linepithema dispertitum | 2 | 0 |
Pheidole spp. | 7 | 6 |
Pogonomyrmex barbatus | 1 | 2 |
Prenolepis imparis | 3 | 0 |
Monomorium ebenium | 1 | 1 |
Discussion
This study found that certain ant species (e.g., C. rubrithorax, L. dispertitum, P. barbatus) exhibited a preference for honey within specific vegetation types when compared to the rest of the ant community. The compensation hypothesis (Davidson, 2005; Kaspari & Yanoviak, 2001) predicts that the attractiveness of a nutrient to an organism is higher the more limiting it is. It is possible that this result is related to the fact that sugar is generally less available on the ground strata than protein (Kaspari & Yanoviak, 2001; Kaspari et al., 2012).
However, when considering the overall community level, no noticeable food preferences were found between the ant communities in the oak forest and the induced grassland vegetation. This lack of preference was evident both in the items carried to their nests and the resources provided on the baits. There were no significant differences in arrival times at the baits between vegetation types. This similarity can be attributed to ants experiencing soil-level heterogeneity similarly in both oak forests and induced grasslands (see below).
The ant communities in both habitats were composed of ant species displaying submissive behavior. This characteristic could be attributed to stress factors such as low temperatures in the oak forest and disturbances in the induced grassland, which potentially reduce ant competition. In the oak forest, low temperatures may favor the presence of cold climate specialists that can forage at low temperatures without the need to display a dominant behavior. In the induced grassland, disturbances may favor the presence of generalist species that take advantage of the absence of dominant species such as the dominant Dolichoderinae (see below). This study emphasizes the significance of ant species’ response to their environment and their adaptability in dealing with disturbances.
Our findings do not provide support for vegetation-scale disparities in food preferences, first and second hypotheses, which proposed greater differences in the food items being carried to the nest in the oak forest and reduced food preferences in the induced grassland. There were no differences between vegetation types in the number of seeds, plants, or arthropods taken by the ant foragers to their nests. Similarly, there was not a preference for a specific food bait (honey or tuna) between the oak forest and induced grassland ant communities. These findings are consistent with previous studies conducted by Radnan et al. (2018) and Castillo-Guevara et al. (2019), which did not identify differences in food preferences or foraging strategies between natural and disturbed vegetation at the community level. However, our observations did reveal that certain ant species exhibited preferences for specific types of food, within particular vegetation types, in accordance with the first hypothesis and second hypothesis but at the species level, which proposed an influence of ant species on food preferences. The specific preferences of these ant species are discussed in detail below.
Table 3
Visit frequency (i.e., number of foragers) per species in each vegetation type and functional groups they belong to (Andersen, 2000). GM = Generalized Myrmicinae, CCS = cold climate specialist, SC = subordinate Camponotini, TCS = tropical climate specialist, O = opportunist.
Subfamily | Specie | Funtional group | Oak forest | Induced grassland |
Myrmicinae | Monomorium ebenium | GM | 753 | 167 |
Myrmicinae | Pheidole sp. 3 | GM | 14 | 236 |
Myrmicinae | Temnothorax sp. | CCS | 5 | 0 |
Myrmicinae | Temnothorax augusti | CCS | 0 | 1 |
Formicinae | Prenolepis imparis | CCS | 289 | 11 |
Formicinae | Camponotus rubrithorax | SC | 0 | 351 |
Formicinae | Nylanderia austroccidua | TCS | 0 | 30 |
Dolichoderinae | Linepithema dispertitum | CCS | 190 | 0 |
Dolichoderinae | Dorymyrmex insanus | O | 0 | 26 |
Pseudomyrmecinae | Pseudomyrmex pallidus | TCS | 0 | 10 |
Total | 1,251 | 832 |
Table 4
Average numerical dominance index (DI) (mean ± SE, n) per ant morphospecies. (-) Unregistered species.
Specie | Oak forest | Induced grassland | ID |
Monomorium ebenium | 0.4 ± 0.1, 13 | 0.3 ± 0.1, 9 | 0.4 ± 0.0, 22 |
Pheidole sp.3 | 0.2 ± 0.2, 4 | 0.2 ± 0.1, 9 | 0.2 ± 0.1, 13 |
Temnothorax sp. | 0 ± 0, 3 | – | 0 ± 0, 3 |
Temnothorax augusti | – | 0, 1 | 0 ± 0, 1 |
Prenolepis imparis | 0.3 ± 0.1, 10 | 0.5 ± 0.5, 2 | 0.3 ± 0.1, 12 |
Camponotus rubrithorax | – | 0.1 ± 0.0, 15 | 0.1 ± 0.0, 15 |
Nylanderia austroccidua | – | 1, 1 | 1, 1 |
Linepithema dispertitum | 0.7 ± 0.1, 3 | – | 0.7 ± 0.1, 3 |
Dorymyrmex insanus | – | 0.2 ± 0.1, 4 | 0.3 ± 0.1, 4 |
Pseudomyrmex pallidus | – | 0 ± 0, 3 | 0 ± 0, 3 |
Pheidole spp., in the induced grassland, were observed carrying a greater quantity of plant elements to their nests in both vegetation types. However, it exhibited a preference for the tuna bait in induced grassland, possibly indicating a supplementary dietary preference. Given the extensive diversity within the Pheidole genus, which encompasses 900 species described worldwide (Wilson, 2003), it is not feasible to categorize them based on a specific food habit and our results suggest that it is omnivorous (Table 6). Nevertheless, these results should be interpreted with caution due to the grouping of species.
Camponotus rubrithorax and P. barbatus, in the induced grassland, as well as L. dispertitum, in the oak forest, were more frequently observed foraging the honey bait compared to other ant species. Camponotus (Mayr) is a genus known for its nectarivorous habits and consumption of other sweet secretions, such as honeydew which coincides with its preference for honey found in the present study (Nettimi & Iyer, 2015). In this study, L. dispertitum predominantly consumed honey, although it is a generalist forager species capable of consuming other types of food as well (Table 5), in the study site it is found exclusively in the oak forest (Cuautle et al., 2016).
Despite Pogonomyrmex (Mayr) has been recognized as a granivorous genus (Pirk & López-de Casenave, 2014), this study provided new insights into the foraging preferences of P. barbatus, revealing a notable inclination towards carbohydrates and plant elements within the induced grassland. Moreover, we even registered individuals transporting arthropods to their nests. These findings (Table 5) strongly indicate that certain species within the Pogonomyrmex genus exhibit a generalist foraging behavior. Other species such as D. insanus, which used resources more intensively in the induced grassland, showed no preference for any of the baits or items taken to the nest, which coincides with the generalist forager behavior, observed in open and disturbed habitats by Cuezzo and Guerrero (2012).
Table 5
Average numerical dominance index (DI) (mean ± SE, n) per ant species in both bait types (tuna, honey). (-) Unregistered species.
Species | Tuna | Honey |
Monomorium ebenium | 0.4 ± 0.1, 18 | 0.4 ± 0.1, 16 |
Pheidole sp.3 | 0.4 ± 0.2, 7 | 0.1 ± 0.1, 8 |
Temnothorax sp. | 0, 1 | 0 ± 0, 2 |
Temnothorax augusti | – | 0, 1 |
Prenolepis imparis | 0.5 ± 0.1, 7 | 0.3 ± 0.1, 10 |
Camponotus rubrithorax | 0.1 ± 0.0, 14 | 0.2 ± 0.0, 12 |
Nylanderia austroccidua | – | 1, 1 |
Linepithema dispertitum | 0.7 ± 0.2, 2 | 0.6 ± 0.3, 3 |
Dorymirmex insanus | 0.6 ± 0.3, 3 | 0 ± 0, 4 |
Pseudomyrmex pallidus | 0 ± 0, 3 | 0, 1 |
Table 6
The biology and ecology of the ant genera found in this study, in the Flor del Bosque State Park (Agosti et al., 2000; AntWiki, n.d.).
Genera | Microhabitat | Food habits |
Camponotus | Ground nesting, decaying wood and in trees | Generalist foragers |
Dorymyrmex | ————– | Generalist foragers |
Linepithema | ————– | Generalist foragers |
Labidus | Epigeous,bivouacs | Predators (Army ants) |
Monomorium | ————– | Generalized foragers, harvesters |
Nylanderia | Nest in leaf litter, soil, or in rotten wood | Generalist foragers |
Pheidole | Soil or decaying wood | Granivores or omnivores |
Pogonomyrmex | Ground nesting | Generalist foragers and granivores |
Prenolepis | ————– | Generalist predators |
Pseudomyrmex | Mostly arboreal (nesters and foragers), few epigaeic | Generalized predators, visit extrafloral nectaries |
Temnothorax | Nesting in ground, and under stones, in wood, and in trees | Generalized foragers and parasites |
The previous results (e.g., C. rubrithorax, Pheidole spp., D. insanus) support the reported food habits of specific ant genera. Nonetheless, it is also possible to interpret these results as ants taking nutrients from the baits that are not currently available or that are limited within their community to supplement their diet (Compensation hypothesis) (Davidson, 2005; Kaspari & Yanoviak, 2001). In this study, significant differences in ant species presence on baits were primarily associated with the presence of ant species on honey baits. This finding aligns with the compensation hypothesis (Davidson, 2005; Kaspari & Yanoviak, 2001), which posits that the utility of a resource remains constant across species and varies solely with availability. The hypothesis predicts a singular limiting resource that is locally in shortest supply. Consequently, habitats with relatively high protein availability should attract ants more inclined towards carbohydrates, and vice versa. Previous research has demonstrated that litter ant communities are limited by carbohydrates, whereas ant arboreal communities face protein limitations in tropical regions (Kaspari & Yanoviak, 2001; Kaspari et al., 2012). Although this study was conducted in a temperate environment, the results might be related to the usual scarcity of sugar in the ground strata compared to protein.
Our results do not align with the second hypothesis, which speculated faster arrival times in the induced grassland. We did not find significant differences in arrival times among vegetation types, ant species, or bait types. Ants could be experiencing similar heterogeneities at the soil level between the oak forest and induced grassland. While it is expected that the presence of more litter in the oak forest could hinder ant movement, within the induced grassland, the vegetation morphology itself (long grass species) could be interfering with ant movement. For instance, Hernández-Flores et al. (2016) observed that the foraging performance of P. barbatus was reduced due to the presence of herbaceous vegetation in plots where regeneration after grazing was permitted.
Dominance hierarchy. Disturbance typically favors the presence of generalist and opportunistic species, hence, we expected to find a lower dominance index within the induced grassland (third hypothesis). Nonetheless, we did not find enough evidence to support this hypothesis as we registered ant communities consisted of submissive species in both habitats. Additionally, the ants did not show a tendency to dominate a specific resource, neither at the bait level nor at the species level, suggesting a lack of food preferences. These results are consistent with the findings of Castillo-Guevara et al. (2019), who registered no significant differences in dominance indices between a native oak forest and an altered agricultural land. Moreover, the authors reported intermediate to low values of dominance within the ant communities of each vegetation type. Our findings could be attributed to the presence of other factors that could have overshadowed the role of competition in shaping the organization of ant communities. For example, a review by Parr and Gibb (2012), which encompassed data from 3 continents, indicated that the trade-off between discovery and dominance occurs primarily when parasitoids are present. In environments without parasitoids, species with high discovery abilities tend to also be dominant (Parr & Gibb, 2012). While our study did not specifically assess the discovery-dominance trade-off, the observed lack of differences in arrival times and similar dominance values among the ants suggest that dominance was not a prominent factor in our study sites. It is possible that factors such as low temperatures in the oak forest, or disturbance in the induced grassland, may have played a role in relaxing dominance. According to Andersen (2000), in disturbed vegetation like the induced grassland, it is expected to observe the presence of subdominant ant species that exploit the absence of dominant species from the native vegetation, such as dominant Dolichoderinae. The findings obtained during this study align with this prediction and provide support for it. Notwithstanding, the absence of ants with high dominance indices in the studied communities does not necessarily indicate a complete lack of dominance hierarchy. In each community, the ants can still be ordered based on their DI. For example, in the oak forest, L. dispertitum had the highest DI value (0.7 DI), M. ebenium had an intermediate value (0.4 DI) and Temnothorax spp. had the lowest value (0.0 DI).
Land use change did not seem to influence food preferences or foraging strategies at the community level. However, we observed an effect at species level, indicating that individual ant species exhibited specific food preferences. Carbohydrates could be the limiting resource in the oak forest and the induced grassland litter ant communities, as some ant species showed preference for honey baits. Although competition is typically considered a key factor in understanding food preferences among ants, it is noteworthy that both, the natural oak forest and induced grassland, were predominantly populated by submissive ant species. Therefore, it appears that other factors instead of competition may be playing a role in shaping food preferences within these communities. In conclusion, our study reveals that ant species may exhibit preferences for specific foods, which could be limited in their environment. The ability of ants to respond to available resources enables them to optimize their nutrient intake, as well as adapt and persist under variable conditions, including disturbances. Understanding the distinct dietary preferences and foraging strategies of ant species within functional groups will provide valuable insights into their ecological roles and potential impacts on ecosystem dynamics. Such investigations would enhance our ability to predict the responses of ants to diverse forms of disturbances in an anthropized world.
Acknowledgements
We appreciate the assistance provided by the authorities of “Flor del Bosque” State Park, coordinator Enrique Martínez Romero (M.S.) and director Mario Alberto Romero Guzmán (MVZ). We would also like to thank Florencio Luna Castellanos for his support with the fieldwork. This study was financed by Consejo Nacional de Humanidades Ciencias y Tecnologías (Conahcyt) as part of a grant awarded to Mariana Cuautle (223033).
References
Abacus Concepts Inc. (1996). Abacus Concepts, Stat View Reference. Berkeley, California.
Agosti, D., Majer, J. D., Alonso, L. E., & Schultz, T. R. (2000). Ants: standard methods for measuring and monitoring bio-
diversity. Washington D.C.: Smithsonian Institution Press.
Andersen, A. (2000). A global ecology of rainforest ants: functional groups in relation to environmental stress and disturbance. In D. Agosti, J. D. Majer, L. E. Alonso, & T. R. Schultz (Eds.), Ants: standard methods for measuring and monitoring biodiversity (pp. 25–34). Washington D.C.: Smithsonian Institution Press.
Andersen, A.N. (1992). Regulation of “momentary” diversity by dominant species in exceptionally rich ant communities of the Australian seasonal tropics. The American Naturalist, 40, 401–420. https://www.journals.uchicago.edu/doi/abs/10.
1086/285419
AntWiki (n.d.). Nylanderia. AntWiki. Consulted 8/20/2024
Arnan, X., Cerdá, X., & Retana, J. (2012). Distinctive life traits and distribution along environmental gradients of dominant and subordinate Mediterranean ant species. Oecologia, 170, 489–500. https://doi.org/10.1007/s00442-012-2315-y
Castillo-Guevara, C., Cuautle, M., Lara, C., & Juárez-Juárez, B. (2019). Effect of agricultural land-use change on ant dominance hierarchy and food preferences in a temperate oak forest. PeerJ, 7, e6255. https://peerj.com/articles/6255/
Cerdá, X., Arnan, X., & Retana, J. (2013). Is competition a significant hallmark of ant (Hymenoptera: Formicidae) ecology? Myrmecological News, 18, 131–147. https://doi.org/10.25849/myrmecol.news_018:131
Cerdá, X., Retana, J., & Cros S. (1997). Thermal disruption of transitive hierarchies in Mediterranean ant communities. Journal of Animal Ecology, 66, 363–374.
Costes-Quijano, R., Meza, A. R., Macías-Juárez, A., Berriel-Mastreta, C. A., Cortés-Atilano, B., Martínez-Romero, L. E. et al. (2006). Plan de manejo Parque Ecológico Recreativo General Lázaro Cárdenas “Flor del Bosque”. [Management Plan of the recreative ecological Park “General Lázaro Cárdenas “Flor del Bosque”]. Ciudad de México: Gobierno del Estado de Puebla/ Secretaría de Medio Ambiente y Recursos Naturales.
Cuautle, M., Vergara, C., & Badano, E. (2016). Comparison of ant community diversity and functional group composition associated to land use change in a seasonally dry oak forest. Neotropical Entomology, 45, 170–9. https://doi.org/
10.1155/2012/516058
Cuezzo, F., & Guerrero, R. J. (2012). The Ant Genus Dorymyrmex Mayr (Hymenoptera: Formicidae: Dolicho-
derinae) in Colombia. Psyche, 51605, 1–24. https://doi.org/
10.1155/2012/516058
Dáttilo, W., Díaz-Castelazo, C., & Rico-Gray, V. (2014). Ant dominance hierarchy determines the nested pattern in ant-plant networks. Biological Journal of the Linnean Society, 113, 405–414. https://doi.org/10.1111/bij.12350.
Davidson, D. W. (2005). Ecological stoichiometry of ants in a New World rain forest. Oecologia, 142, 221–231. https://doi.org/10.1007/s00442-004-1722-0
Dejean, A., & Corbara, B. (2003). A review of mosaics of dominant ants in rainforests and plantations. In Y. Basset, V. Novotny, S. E. Miller, & R. L. Kitching (Eds.), Arthropods of tropical forests: spatio-temporal dynamics and resource use in the canopy (pp 341–347). Cambridge: Cambridge University Press.
Ellis, E. C., Goldewijk, K. K., Siebert S., Lightman, D., & Ramankutty, N. (2010). Anthropogenic transformation of
the biomes, 1700 to 2000. Global Ecology and Biogeo-
graphy, 19, 589–606. https://doi.org/10.1111/j.1466-8238.20
10.00540.x
Fellers, J. H. (1987). Interference and exploitation in a guild of Woodland ants. Ecology, 68, 1466–1478. https://doi.org/
10.2307/1939230
Fellers, J. H. (1989). Daily and seasonal activity in woodland
ants. Oecologia, 78, 69–76. https://doi.org/10.1007/BF0037
7199
Foley, J. A., DeFries, R., Asner, G. P., Barford C., Bonan, G., Carpenter, S. R. et al. (2005). Global consequences of land use. Science, 309, 570–574. https://doi/10.1126/science.1111772
Hoffmann, B. D., & Andersen, A. N. (2003). Responses of ants to disturbance in Australia, with particular reference to functional groups. Austral Ecologyl, 28, 444–464. –https://doi.org/10.1046/j.1442-9993.2003.01301.x
Houdria, M., Salas-López, A., Orivel. J., Bluthgen, N., & Menzel, F. (2015). Dietary and temporal niche differentiation in tropical ants – can they explain local ant coexistence? Biotropica, 47,208–217. https://doi.org/10.1111/btp.12184
Hernández-Flores, J., Osorio-Beristain. M., & Martínez-Garza. C. (2016). Ant foraging as an indicator of tropical dry forest restoration. Environmental Entomology, 45, 991–994. https://doi.org/10.1093/ee/nvw054
Kaspari, M., Donoso, D., Lucas, J. A., Zumbusch, T., & Kay, A. D. (2012). Using nutritional ecology to predict community structure: a field test in Neotropical ants. Ecosphere, 3,93. https://doi.org/10.1890/ES12-00136.1
Kaspari, M., & Yanoviak, S. P. (2001). Bait use in tropical litter and canopy ants-evidence of differences in nutrient limitation. Biotropica, 33,207–211. https://doi.org/10.1646/
0006-3606(2001)033[0207:BUITLA]2.0.CO;2
Łaska, G. (2001). The disturbance and vegetation dynamics: a review and an alternative framework. Plant Ecology, 157, 77–99. https://doi.org/10.1023/A:1013760320805
Lynch, J. F., Balinsky, E. C., & Vail, S. G. (1980). Foraging patterns in three sympatric forest ant species, Prenolepis imparis, Paratrechina melanderi and Aphaenogaster rudis (Hymenoptera: Formicidae). Ecological Entomology, 5,353–371. https://doi.org/10.1111/j.1365-2311.1980.tb01160.x
Mackay, W., & Mackay, E. (1989). Clave de los géneros de hormigas en México (Hymenoptera: Formicidae). El Paso, Texas: The University of Texas.
Nettimi, R. P., & Iyer, P. (2015). Patch fidelity in Camponotus compressus ants foraging on honeydew secreted by treehoppers. Current Science, 109, 362–366.
Parr, C. L. (2008). Dominant ants can control assemblage species richness in a South Africa savanna. Journal of Animal Ecology, 77, 1191–1198. https://doi.org/10.1111/j.13
65-2656.2008.01450.x
Parr, L., & Gibb, H. (2012). The discovery-dominance trade-off
is the exception, rather than the rule Journal of Animal
Ecology, 81, 233–241. https://doi.org/10.1111/j.1365-2656.
2011.01899.x
Pirk, G. I., & López-de Casenave, J. (2014). Effect of harvester ants of the genus Pogonomyrmex on the soil seed bank around their nests in the central Monte desert, Argentina. Ecological Entomology, 39, 610–619. https://doi.org/10.1111/een.12140
R Core Team (2022). R: a language and environment for statistical computing. Vienna, Austria: R Foundation
for Statistical Computing. https://www.R-project.org/
Radnan, G. N., Gibb, H., & Eldridge, D. J. (2018). Soil surface complexity has a larger effect on food exploitation by ants than a change from grassland to shrubland. Ecological Entomology, 43, 379–388. https://doi.org/10.1111/een.12510
Rico-Gray, V., & Oliveira, P. S. (2007). The ecology and evolution of ant-plant interactions. Chicago: University of Chicago Press.
Sala, O. E., Chapin, F. S., Armesto, J. J., Berlow, E., Bloomfield, J., Dirzo. R. et al. (2000). Global biodiversity scenarios for the year 2100. Science, 287,1770–1774. https://doi.org/10.1126/science.287.5459.17
Santini, G., Tucci, L., Ottonetti, L., & Frizzi, L. (2007). Competition trade-offs in the organisation of a Medite-
rranean ant assemblage. Ecological Entomology, 32,
319–326. https://doi.org/10.1111/j.1365-2311.2007.00882.x
Savolainen, R., Vepsäläinen, K., & Wuorenrinne, H. (1989). Ant assemblages in the taiga biome: testing the role of territorial wood ants. Oecologia, 81, 481–486. https://doi.org/10.1007/BF00378955