Taxonomic and Functional Diversity of Bees in Traditional Agroecosystems and Tropical Forest Patches on the Yucatan Peninsula

Abstract

Background and Research Objectives: Habitat matrices of intensive agricultural use are generally inhospitable to native bees in fragmented forests. However, in some tropical regions of the world, agricultural landscapes are dominated by traditional agroecosystems, which harbor high plant diversity and are subject to low-intensity management. These agroecosystems can therefore provide suitable habitats and important floral resources for the bee community. The objective of this study was to compare the taxonomic and functional diversity of bees in traditional agroecosystems and forest patches within an agricultural landscape of the Yucatan Peninsula. Methods: Sampling was conducted in two traditional agroecosystems (homegardens and a rainfed polyculture known as milpa) and forest patches as a control (N= 24 sites in total. Hereafter: habitats), using two complementary sampling techniques: pan traps and a sweep net. Taxonomic and functional diversity metrics were calculated and compared among habitats. Results: The three habitats were generally similar in terms of taxonomic and functional diversity. Differences were only detected among habitats in the inverse of Simpson’s diversity index and number of functional groups (functional entities), with higher values in the agroecosystems than in the forest. Conclusion: Taxonomic and functional diversity was similar in the traditional agroecosystems and forest patches, suggesting that these agroecosystems can provide temporal adequate resources for most bee functional groups and that movement of bees is possible among these habitats. Conservation Implications: Maintenance of low-intensity management practices and high (agro)biodiversity found in traditional agroecosystems is crucial for the conservation of native bees. It is therefore important to incorporate these systems into management strategies at the landscape level. Since traditional agroecosystems and forests have different land tenure systems (government, private, and communal), conservation strategies at this level require the involvement of different sectors of society.

Keywords

agricultural landscapes homegardens landscape connectivity milpa wild bees

Introduction

The ongoing accelerated change and increasing intensification of land use for agricultural purposes have led to severe forest fragmentation and habitat loss, anthropogenic processes that are also identified among the main drivers of biodiversity loss worldwide (Bełcik et al., 2020; Bonfim et al., 2021). Tropical forests are particularly vulnerable to these processes, as some critical ecosystem functions (e.g. pollination and seed dispersal) have little functional redundancy; i.e. they rely heavily on specialized biotic interactions (Andresen et al., 2018; Zhang & Zang, 2021; Török et al., 2022). This accelerated conversion of forest areas to cropland therefore raises the need to monitor the status of biodiversity (including agrobiodiversity) to establish effective conservation strategies (Andresen et al., 2018; Brito et al., 2018). The current dominant landscape in the tropics is a mosaic of cultivated fields with different types and intensities of management, and forest patches at different successional stages (Carrié et al., 2017). The distribution and permanence of native fauna in these mosaics depend on the characteristics of the landscape units, e.g. the type and intensity of their land use, their suitability as a habitat, and the connectivity within the landscape, as well as the attributes of the species present, such as their ability to move, degree of specialization in resource use (food and shelter), or sensitivity to stressors in agricultural habitats (Rader et al., 2014; Pinto et al., 2020; Millard et al., 2021).

Native bees form a taxonomic group widely recognized for its value as an indicator of habitat disturbance due to its high taxonomic diversity and variety of functional groups (Mélendez-Ramírez et al., 2014; Twerd & Sobieraj-Betlinska, 2020). Bees also play a vital role in highly important ecosystem functions such as the pollination of wild and cultivated plants, and are therefore also involved in ecosystem services of provision (Klein et al., 2007; Basu et al., 2016). When forest fragmentation occurs, it reduces the size and increases the isolation of habitat patches, leading to a reduction in floral resources and potential nesting sites for the bees, as well as impacting the richness and/or abundance of certain species (e.g. Meneses-Calvillo et al., 2010; Brito et al., 2018; Rahimi et al., 2022). The response of bees to anthropogenic habitat loss and transformation depends on certain bee characteristics. For example, body size is positively correlated with flight distance and foraging time, and large bees may also be less susceptible to habitat reduction and isolation (Williams et al., 2010; Cohen et al., 2022). Chronic soil disturbance can also reduce the number of suitable nesting sites, making intensively managed areas unfavorable for ground-nesting species (Odanaka & Rehan, 2019; Kohler et al., 2020). The use of agrochemicals can also decrease floral visitation rates in wild bees (i.e. via bee intoxication and/or a reduction in floral resources), particularly of solitary and oligolectic species given their high degree of specialization and reduced mobility (Bloom et al., 2021; Siviter et al., 2021). However, when habitat transformation increases heterogeneity at a meso(intermediate)- or large-scale, this can also diversify the supply of resources and thus indirectly favor the permanence of a more diverse group of bees (Odanaka & Rehan, 2019; Rodríguez et al., 2021; Peters et al., 2022).

Studies on the effect of different anthropogenic processes on bee communities in agricultural landscapes have mainly been conducted in temperate habitats (e.g. Kennedy et al., 2013; Bentrup et al., 2019). On the other hand, there have been relatively fewer studies in tropical regions, and those that do exist have focused on landscapes dominated by intensive monocultures or semi-intensive, poorly diversified polycultures (IPBES, 2016; Escobedo-Kenefic et al., 2020; Vogel et al., 2023). In contrast to most temperate agricultural landscapes, the heterogeneity in land use type and intensity of tropical agricultural landscapes is typically much higher (Basu et al., 2016; Laha et al., 2020; Escobedo-Kenefic et al., 2022). This heterogeneity may provide complementary habitats that together could offer a greater variety of food resources to bees than a landscape dominated by intensive monocultures (Hotspot et al., 2008; Escobedo-Kenefic et al., 2022). One particular case, still common in some regions of the tropics, is that of agricultural landscapes dominated by traditional agroecosystems (e.g. homegardens and other diversified agroforestry systems; Perfecto & Vandermeer, 2008) mixed with forest remnants. These agroecosystems present a high native plant diversity, forest-like vertical structures, and active agroecological management, characteristics that make them a suitable habitat for some native species, while also functioning as stepping stones for species with a preference for the forest interior (e.g. Galluzi et al., 2010; Rajagopal et al., 2021; Villicaña-Hernández et al., 2020).

The agricultural landscape of the Yucatan Peninsula has been characterized since ancient times by the presence of milpa (a rainfed polyculture dominated by maize, beans, and squash), homegardens dominated by trees with fleshy fruits, and forest remnants at different successional stages (Lope-Alzina & Howard, 2012; Nigh & Diemont, 2013; Ford & Topsey, 2019). Together, these habitat units provide diverse and complementary resources that are fundamental for the subsistence of the indigenous people who inhabit this region (Rodríguez-Robayo et al., 2020; Serralta-Batun et al., 2024). On the Yucatan Peninsula, previous studies conducted in milpas have found that bee richness and abundance are higher compared to intensive monocultures (Briggs et al., 2013; Landaverde et al., 2017; Vides-Borrell et al., 2019). However, in contrast to other agroecosystems and forest remnants, the suitability of homegardens as a habitat and their value in terms of bee conservation remain to be determined. The limited studies on bee diversity in traditional agroecosystems have also focused on describing the taxonomic diversity of bees (Meléndez-Ramírez et al., 2002; Meléndez-Ramírez et al., 2016 Landaverde et al., 2017; Vides-Borrell et al., 2019). Although an analysis of taxonomic diversity can indicate how species richness and abundance change in different habitat units of the landscape, this approach provides limited information on the environmental filters operating in each habitat type and on the functional groups affected, aspects that could be addressed with an analysis of functional diversity (Díaz & Cabido, 2001; Villegér et al., 2008; Mouillot et al., 2014).

This study addressed the taxonomic and functional diversity of native bees in the main landscape units (hereafter: habitats) of the agricultural landscape of the Yucatan Peninsula: the milpa, homegarden, and forest. The taxonomic composition and distribution of some functional traits of the prevalent bee species in these habitats were investigated. The habitats differ in terms of plant species richness and level of management intensity (Serralta-Batun et al., 2024), with the most diverse and least intensively managed habitat being the forest, and the least diverse and most intensively managed being the milpa. Homegardens occupy an intermediate position between these two variables (Serralta-Batun et al., 2024). Our specific research questions were: Are there differences in native bee taxonomic and functional diversity among the three dominant habitats in the landscape? and which functional traits are favored by the prevalent environmental filters in these habitats? We predicted that there would be greater taxonomic and functional diversity in the forest than in the milpa, and an intermediate level would be present in the homegardens, given the gradient of management intensity and plant diversity observed. In terms of functional groups, generalist (in terms of diet and nest site) and large bee species were expected to predominate in the less diverse and more intensively-managed habitat (milpa) while functional groups of bees with more specialized requirements and lower capacity for movement would be found mainly in the more diversified and less intensively-managed habitats (homegarden and forest).

Methods

Study System

The study area is located in the south of the state of Yucatan, Mexico, and includes the municipalities of Tahdziú and Tzucacab (20°00'-13' N, 88°-89°02’W; Figure 1). The climate of the area is warm sub-humid with rainfall in summer and the mean annual temperature is 25.8 °C (Camacho-Villa et al., 2021). The numbers of inhabitants were 15,346 and 5,854 in Tzucacab and Tahdziú, respectively (Instituto Nacional de Estadística y Geografía [INEGI], 2020). In both municipalities, the population is mostly of Maya ethnicity (INEGI, 2020). The livelihood of most families is based on primary activities such as agriculture, mainly milpa (Serralta-Batun et al., 2024), complemented with food resources produced in their homegardens (fruits, vegetables, and animal protein), as well as the extraction of timber (e.g. Psicidia piscipula, Diospyros cuneata) and non-timber (e.g. Sabal yapa, Sabal mexicana) forest resources from the tropical forest remnants (Cruz-Cortés et al., 2019).

Figure 1.

Location of the study area and habitats sampled: Forest patches (forest), homegardens and milpa. The upper right insert shows the Yucatan Peninsula, Mexico.

In contrast to traditional itinerant milpa, the current milpa in the study area is sedentary (cultivated in the same place every year), rainfed, and has an average area of 1.57 ± 0.28 ha (n= 8; hereafter: mean ± 1 SE), a plant species richness of 35 ± 4.2 species, and a Shannon-Wiener diversity (H´) of 2.9. The dominant species in the milpa are Zea mays, Phaseolus vulgaris, and various species of Cucurbita. On the other hand, the homegardens in the area present a richness of 71 ± 1.4 plant species and H´= 3.8. The average area of the sampled homegardens was 0.8 ± 0.22 ha (n= 8). The dominant species in this agroecosystem were Citrus aurantium, C. sinensis, Capsicum chinense, and Cnidoscolus aconitifolius (Serralta-Batun et al., 2024).

The traditional agroecosystems (homegarden and milpa) in the region typically do not have irrigation systems. Irrigation of the homegardens is manual, selective (for some species only), and conducted during the dry season. In the case of the milpa, it is a rain-fed system, which means that it is dependent on the occurrence and quantity of rainfall. There are initiatives to maintain the use of agroecological practices in the region, such as the application of organic fertilizers and pest management with bioinsecticides (Serralta-Batun et al., 2024). Forty percent of the study milpas did not use any herbicides and the remaining 60% had only used herbicides on one occasion over the last three years.

The milpa and the homegardens are both surrounded by a secondary tropical sub-deciduous forest at different stages of succession. In the region, a woody plant richness of between 42 and 65 species has been reported (Rico-Gray & García-Franco, 1992), with an H´ index of 3.55-.82, and a dominance of the species Bursera simaruba, Croton reflexifolius, Diospyros cuneata, Gymnanthes lucida, and Lysiloma latisiliquum (Zamora-Crescencio et al., 2008). Some extractive activity is maintained in the forest patches to obtain firewood or construction materials (Serralta-Batun et al., 2024). Some wild plant species and wild relatives of cultivated plants are shared between the homegardens and forest remnants (e.g. L. latisiliquum; Cordia dodecandra; wild Carica papaya and wild Cnidoscolus aconitifolius), while only a few species are shared between forest and milpa (wild Capsicum annum; wild Cnidoscolus aconitifolius) since most plants cultivated in the latter are fully domesticated and thus rely on human assistance for their reproduction.

Sampling

The study was conducted during the time that the milpa is cultivated (May 2021 and January 2022, a period of nine months), which is when the three habitat types considered in this research are found simultaneously, and coincided with the rainy season in the study area. The main factor of interest was habitat type, with three levels: forest patches, homegardens, and milpas. Eight sites per habitat type were selected, for a total number of 24 sites. This sample size is similar to those in previous studies on the topic (e.g. Rader et al., 2014; Forrest et al., 2015; Vides-Borrell et al., 2019). In the case of the secondary forest, site selection was random. The selected patches had between 8 and 15 years of regeneration, a successional stage locally known as Ka'anal hubché (“tangle of tall trees” in Maya; González-Cruz et al., 2015). In the case of the agroecosystems, selection depended on authorization from the landowner. Therefore, in the case of no authorization, the closest alternative site with authorization was selected. The average distance between the 24 selected sites was 5.96 ± 1.70 km (range 4.63-9.33 km), which exceeds the average distance (2.7 km) that some bees can fly in a single day (Greenleaf et al., 2007).

Native bees were sampled in the 24 sites (eight per habitat type) with three-bowl pan traps: one blue bowl, one white bowl, and one yellow bowl, colors that mimicked those of the flowers of the dominant species in the three habitat types (Parys et al., 2020). Nine such traps were established in each of the 24 sites (8 sites x 3 habitats x 9 traps= 216 traps in total), at a height of 50 cm above ground level and 5 m apart. Each bowl was filled with 150 mL of water and 2 mL of liquid detergent as a surfactant, and each trap was left for 7 h (08:00-15:00 h). This procedure was repeated approximately bimonthly (every 8-9 weeks) for each habitat during the nine months in which the study was conducted (four sampling periods during the study). Thus, the pan traps were left in each site for 28 h in total (7 h x 4 periods). The time elapsed to complete the sampling of the 24 sites was 10-14 days within each sampling period. Although the total area sampled (400 m²) was kept constant in all habitat types, the distance between traps and their spatial arrangement varied slightly due to differences in vegetation distribution and plant density among the habitats. In the milpa and homegardens, the traps were distributed in a 20 x 20 m (400 m²) quadrat with a distance of 3-4 m between traps, while in the forest, two 20 x 1 m strip transects (2 transects x 20 m² = 400 m² sampled) were established at a distance of 20 m apart.

In addition to the sampling technique with pan traps, complementary sampling was conducted with a sweep net for two hours at a distance of at least 30 m from the traps to avoid interference. Both sampling techniques are suitable for forest and agroecosystems (Prado et al., 2017). As with the pan traps, sweep netting was conducted at a frequency of once every ca. two months per site for the total duration of the study (2 h x 24 sites x 4 times = 192 h of sampling in total). To reduce the border effect, the traps were set and netting was conducted in the central area of the quadrants/transects. The bee specimens collected (using both sampling techniques) were preserved in 70% alcohol and subsequently identified to the lowest taxonomic level possible. When this manuscript was submitted, the collected and identified specimens were in quarantine, pending admission to the Regional Entomological Collection of the Autonomous University of Yucatan (CER-UADY, by its Spanish acronym). The collection sites were not located within a protected natural area, nor were they found within the distribution range of any species on the IUCN red list; consequently, a collection permit was not required.

Data Analyses

Once the collected bee specimens were identified to the finest level possible, the Hill numbers, or effective number of species of order 0 (D⁰), order one (D¹), and order two (D²), were calculated (Jost, 2006). We chose these metrics because they have intuitive mathematical properties, are easily interpreted, and are highly comparable since they feature commonly in the literature. D⁰ corresponds to species richness, while D¹ weights species by their frequency, without favoring rare or common species, and its relationship to other common diversity indices is known (D¹= exp H'), and D² is influenced by abundant and common species and is also the inverse of Simpson's diversity index (Jost, 2006). The expected richness was calculated with the Chao1 estimator. All these metrics were calculated with the iNEXT package version 3.0 (Hsieh et al., 2016) and were compared among habitats with a Chi-squared test of homogeneity, which tested the alternative hypothesis that the observed values of the metrics will differ from that expected by chance. Since our study was focused on native bees, Apis mellifera was excluded from this and all data analyses.

To identify differences among habitats in terms of bee species composition, an ordination analysis was performed using non-parametric multidimensional scaling (NMDS) based on the Bray-Curtis dissimilarity index. To define the degree of clustering, the centroids of each habitat were identified using the convex hull approach and ellipses were plotted with 95% confidence for each habitat type. Likewise, species composition was compared among the three habitat types and between the two municipalities (Tahdziú vs. Tzucacab) with a permutational analysis of variance using the Bray-Curtis distance matrix (PERMANOVA) with 999 permutations. The municipality was also considered as a block in the model; therefore, permutations were performed only within the same municipality using the argument “strata” from the function “adonis” of the vegan package (Oksanen et al., 2012) for R 4.0.3 (R Core Team, 2022). Six categorical functional traits related to disturbance tolerance, habitat use, and life history were used to characterize the functional diversity of native bee communities across the habitat types. These traits were (i) nesting behavior, (ii) sociability, (iii) diet, (iv) size, (v) hairiness, and (vi) pollen-collecting structure (sensu Michener, 2007; see Table 1 for definition of trait states). In particular, nesting location is related to the preferred sites for nesting, a critical issue for bee population establishment (Forrest et al., 2015). The level of sociability is linked to shorter (solitary) or longer (eusocial) seasonal activity and may therefore affect the temporal patterns of bee activity (Kratschmer et al., 2019). The diet of the bee species reflects the type of resources to which they have access, species with more specialized diets tend to be more susceptible to disturbance (Michener, 2007). Size is related to home range and the distance over which a female can fly and collect pollen and nectar. Larger species would be expected to utilize a wider diversity of habitats and resources (Greenleaf et al., 2007). The density and distribution of hairiness reflect the plant species being visited (i.e. bees carry the pollen of only those plants in which the anther position matches their hair distribution) and the quantity of pollen transported by the bees (the amount of hair is positively related to the quantity of pollen) (Stavert et al., 2016). Finally, pollen transport structure is also an indicator of pollinator or foraging efficiency being correlated with the ability to carry more pollen per trip (i.e. bees with larger corbiculae transport more pollen; Michener, 2007; Coutinho et al., 2021).

Table 1.

Functional traits and trait categories (trait state) of the bees used in this study. Each category is described in the last column.

Functional trait	Trait state	Description
Nesting location	Above ground Below ground Parasitic	Nesting above soil level Nesting below soil level Nest invasion
Sociability	Social Solitary FVL Parasitic	Bees living in colonies. Bee constructs her own nest and provides food for her offspring. Bees with variable lifestyle (solitary, subsocial, semisocial, quasisocial, primitive eusocial, and communal) Bee enters the nest of a host and lays an egg in a cell
Diet specialization	Oligolectic Polylectic Parasitic	Pollen specialist Pollen generalist Females do not collect pollen
Size	Very small Small Medium, Large Huge	< 5 mm 6 – 8 mm 9 – 11 mm 12 – 15 mm > 15 mm
Hairiness	Dense Sparse	Hair restricted to one part of the body Hair on several parts of the body
Pollen structure	Corbiculae Ventral scopa Scopa1 Scopa2 No structure	Part of the tibia on the hind legs of the female Hairs on the ventral side on the metasoma Hairs restricted to hind tibia and basitarsus Hairs on femur, trochanter, hind coxae, and middle No structure for pollen collection

Once the bees were characterized within different trait states, frequency distributions were compared among habitat types for each trait with a Chi-square test of homogeneity, in which the null hypothesis was that the frequency distribution per functional trait category will be equal across the three habitat types.

Since all bee functional traits considered in this study were categorical, a first approach to functional diversity was made using an analysis based on functional entities (FE), which are categories defined from the unique combinations of all of the traits (Mouillot et al., 2014). From the frequency distribution analysis of these FE, relevant aspects of the community can be inferred, including its level of redundancy vs. functional vulnerability (Mouillot et al., 2014). The metrics calculated from the FE were (i) functional redundancy (FRed), defined as the average number of species per FE, (ii) functional over-redundancy (FOR), defined as the percentage of FE with more species than the average of all FE, and (iii) functional vulnerability (FVuln), which is the percentage of FE with a single species (Mouillot et al., 2014). These metrics were calculated with the mFD package of R, 4.0.3 (Magneville et al., 2022), and compared among habitats with a non-parametric Kruskal-Wallis test using the same package and version.

As a complement to the FE-based analysis, a series of functional metrics were calculated that collectively describe the species distribution in the multivariate functional space using the convex hull approach generated from a Gower distance matrix based on the characters described above (Villéger et al., 2008). The multivariate functional metrics calculated were (i) functional richness (FRic), which represents the functional space occupied by the species (Mason et al., 2005), (ii) functional evenness (FEve), which indicates the regularity with which species are distributed in this space, (iii) functional divergence (FDiv), which is a measure of functional similarity among the dominant species in a community, where high values are associated with a high degree of niche differentiation between dominant species, which could act to reduce competition and increase the magnitude of ecosystem processes as a result of more efficient resource use (Mason et al., 2005; Mouchet et al., 2010), and finally (iv) functional dispersion (FDis), which defines how far the most abundant species move away from the center of the functional space (Banaszak-Cibicka & Dylewski, 2021). The formulae for calculating these metrics were those outlined by Mason et al. (2005), Villéger et al. (2008), and Laliberté & Legendre (2010), which are presented in the online Appendix S1. These metrics were calculated for each site and compared among habitats (fixed effect) using a linear mixed- effects model with the municipality as a random effect. The multivariate functional diversity metrics and linear mixed-effect model were calculated with the FD package and lme4 packages, for R, version 4.0.3, respectively. The raw data from this study are available as supplementary material in Appendix S3.

Results

Taxonomic Diversity

A total number of 451 individual bees were recorded, belonging to 50 species across the entire landscape. The family Apidae was the best represented in terms of species richness (48.2% of the species recorded), followed by the family Halictidae (46.3%) and, in third place with much lower richness, the family Megachilidae (1.7%). The Apidae species with the highest relative abundance were: Nannotrigona perilampoides (7.7%), Trigona fulviventris (7.3%), and Peponapis limitaris (6.4%). While in the family Halictidae these were: Lasioglossum sp2 (7.7%), Lasioglossum sp1 (7%), and Augochloropsis metallica (6.8%).

Neither the observed nor expected species richness differed significantly in the three habitats: forest (D⁰ = 21, expected richness = 30), homegarden (D⁰= 35, expected richness = 51), and milpa (D⁰= 39, expected richness = 47) (Table 2). D¹ also did not differ significantly among the three habitats (Table 2). However, the inverse of Simpson’s index was higher in the two traditional agroecosystems (homegarden = 16.41, milpa = 19.41) than in the forest (5.80) (Table 2). The species with the highest relative abundance within each habitat were as follows: N. perilampoides (37.3%), T. fulviventris (11.9%), and P. utahensis (5.9%) in the forest; A. metallica (11.2%), Lasioglossum sp1 (10.5%), and P. bilineata (7.9%) in the homegardens; and finally, Lasioglossum sp2 (9.4%), A. nygrocyanea (8.1%), and P. limitaris (8.1%) in the milpa.

Table 2.

Observed (D⁰) and expected species richness of bees, true diversity of order 1 (D¹) and of order 2 (D²; Inverse of Simpson’s diversity index) in tropical forest patches (Forest) and two traditional agroecosystems: Homegarden and Milpa on the Yucatan Peninsula, Mexico. The results of the statistics are shown in the last column. Different superscript letters indicate statistically significant differences between habitat pairs.

Habitat
Metric	Forest	Homegarden	Milpa	Statistics
D⁰	21	35	39	χ² = 5.64
Expected richness	30	51	47	χ² = 5.82
D¹	10.8	22.1	24.6	χ² = 5.57
D²	5.8^a	16.3^b	19.4^b	χ² = 7.35*

*p < 0.05.

Considering bee species composition, there was greater similarity between the two agroecosystems (Bray-Curtis dissimilarity distance x -1 = 0.66) than between any of the agroecosystems and the forest (homegarden vs. forest = 0.33; milpa vs. forest = 0.31). The three habitats shared 16 bee species, prominent among which for their relative abundance were the following species: Lasioglossum sp2 (7.7%), N. perilampoides (7.7%), T. fulviventris (7.3%), and Lasioglossum sp1 (7.0%). In addition to these species, the milpa and the homegarden shared 12 (28 species shared in total), including P. limitaris (6.4%), Lasioglossum sp3 (1.5%), Ancyloscelis sp1 (1.3%), and Exomalopsis mellipes (1.1%). Similarly, the forest and the milpa shared one bee species (Eulaema polychrome, 0.4%) as well as the 16 species shared among all three habitats (17 species shared in total). The homegardens and forest were the habitat pair with the fewest shared species, the identities of which were the same as the 16 bee species shared between the three habitats (see online Appendix S2). Four species were exclusive to the forest, seven species were exclusive to the homegardens, and 10 were exclusive to the milpa. The identities of the exclusive species are presented in Table 3.

Table 3.

Exclusive bee species and their relative abundance (values in brackets) found in forest patches (Forest), homegardens, and milpas in an agricultural landscape of the Yucatan Peninsula, Mexico. Different morphospecies within the same genus are identified with consecutive numbers (e.g. sp1, sp2…sp n).

Habitat
Forest	Homegarden	Milpa
Anthidium macusolum (0.2)	Ceratina sp6 (0.6)	Augochlora sp3 (0.6)
Exomalopsis sp1 (0.2)	Augochlora neglectula (0.2)	Augochlora smaradigma (0.4%)
Hylaeus sp1 (0.2)	Coelioxys sp1 (0.2)	Augochlora sp4 (0.4)
Ceratina sp5 (0.2)	Gaesischia sp1 (0.2)	Peponapis sp1 (0.4)
	Lasioglossum sp (0.2)	Centris trigonoides (0.2)
	Megachile albitarsis (0.2)	Megachile sp1 (0.2)
	Megachile habilis (0.2)	Megachile sp2 (0.2)
		Megachile sp3 (0.2)
		Paratetrapedia moesta (0.2)
		Xylocopa muscaria (0.2)

The NMDS demonstrated a considerable overlap among the three habitats in terms of species composition (Figure 2). This result agreed with that of the PERMANOVA, where no significant differences were found among habitat types (F₂, ₁₆ = 0.36, P = 0.13; Figure 2) or between municipalities (F₁,₁₆ = 0.57, P = 0.11).

Figure 2.

Ordination plot of bee species composition in forest patches, home gardens and milpas (ellipses) in an agricultural landscape of Yucatan, Mexico. The dimensions of the axes were obtained from non-metric multidimensional scaling based on the Bray-Curtis dissimilarity index. Ancsp1 = Ancyloscelis sp1; Antmacu = Anthidium macusolum; Augauri = Augochlora aurífera; Augnyg = Augochlora nygrocyanea; Augsma = Augochlora smaradigma; Augsp = Augochlora sp; Augsp1 = Augochlora sp1; Augsp2 = Augochlora sp2; Augsp3 = Augochlora sp3; Augsp4 = Augochlora sp4; Augneg = Augochlorella neglectula; Augmet = Augochloropsis metallica; Centrig = Centris trigonoides; Cerexi = Ceratina eximia; Cernau = Ceratina nautlana; Cersp2 = Ceratina sp2; Cersp3 = Ceratina sp3; Cersp5 = Ceratina sp5; Cersp6 = Ceratina sp6; Coesp1 = Coelioxys sp1; Eugvir = Euglossa viridissima; Eulpol = Eulaema polychroma; Exomel = Exomalopsis mellipes; Exosp1 = Exomalopsis sp1; Frinig = Frieseomelitta nigra; Gaesp1 = Gaesischia sp1; Hylsp1 = Hylaeus sp1; Lasisp = Lasioglossum sp; Lassp1 = Lasioglossum sp1; Lassp2 = Lasioglossum sp2; Lassp3 = Lasioglossum sp3; Lassp4 = Lasioglossum sp4; Megalb = Megachile albitarsis; Megha = Megachile habilis; Megsp1 = Megachile sp1; Megsp2 = Megachile sp2; Megsp3 = Megachile sp3; Nanper = Nannotrigona perilampoides; Parmoe = Paratetrapedia moesta; Parbil = Partamona bilineata; Pepli = Peponapis limitaris; Pepsmi = Peponapis smithi; Pepsp1 = Peponapis sp1; Peput = Peponapis utahensis; Plefro = Plebeia frontalis; Scapec = Scaptotrigona pectoralis; Triful = Trigona fulviventris; Xylmus = Xylocopa muscaria; Xylona = Xylocopa nautlana.

Functional Diversity

Overall, across all habitat types, the proportion of above- (relative frequency= 0.49) and below-ground (0.51) nesting bee species was very similar. The variable lifestyle form was the most predominant (0.45), followed by solitary species (0.29) and social bees (0.24). Regarding diet breadth, polylectic feeding species (0.86) were more frequent than oligolectic species (0.14). In terms of size and hairiness, small-sized (0.41) and sparsely hairy bees (0.70) were predominant. Regarding pollen-carrying structures, the order of frequency were: bees with a scopa 2 (0.64), corbiculate bees (0.23), and species with a scopa 1 (0.13) (Table 4). However, the observed distribution of bee species frequencies among functional states did not differ statistically among habitats for any trait (Table 4; Appendix S2).

Table 4.

Percentage of bee species (frequency shown in brackets) grouped into categories of functional traits for forest patches, homegardens and milpas in an agricultural landscape of the Yucatan Peninsula, Mexico. The results of statistical comparisons among habitats for each trait are shown in the last column, p > 0.05 in all cases.

Trait	Trait state	Forest % (frequency)	Homegarden % (frequency)	Milpa % (frequency)	Statistic
Nesting location	Above ground	43 (9)	51 (20)	49 (20)	χ² = 2.12
	Below ground	57 (12)	46 (18)	51 (21)
	Parasitic	0 (0)	3 (1)	0 (0)
Sociability	Social	32 (7)	24 (9)	22 (9)	χ² = 2.55
	Solitary	27 (6)	29 (11)	32 (13)
	FVL	41 (9)	45 (17)	49 (20)
	Parasitic	0 (0)	2 (1)	0 (0)
Diet	Polylectic	90 (19)	83 (30)	87 (34)	χ² = 0.27
	Oligolectic	10 (2)	14 (5)	13 (5)	χ² = 0.27
	Parasitic	0 (0)	3 (1)	0 (0)
Size	Vey small	9.5 (2)	6 (2)	5 (2)	χ² = 3.63
	Small	38 (8)	46 (15)	38 (15)
	Medium	34 (7)	21 (7)	21 (8)
	Large	14 (3)	24 (8)	28 (11)
	Huge	4.5 (1)	3 (1)	8 (3)
Hairiness	Dense	19 (4)	29 (10)	36 (14)	χ² = 1.88
Hairiness	Sparse	81 (17)	71 (25)	64 (25)	χ² = 1.88
Pollen collection structure	Corbiculae	35 (7)	25 (9)	28 (11)	χ² = 4.12
	Ventral scopa	0 (0)	8 (3)	8 (3)
	Scopa 1	20 (4)	14 (5)	18 (7)
	Scopa 2	45 (9)	50 (18)	46 (18)
	No structure	0 (0)	3 (1)	0 (0)

From the combination of the different functional trait states, 24 FE were obtained. The number of FE differed significantly among habitats, with a higher average number of FE in the milpa (9 ± 2.39) and homegarden (6.62 ± 2.13) than in the forest (5.25 ± 1.25; Table 5). The functional redundancy showed that the FE consisted of slightly more than one species in the agroecosystems (homegarden = 1.45 ± 0.37, milpa = 1.58 ± 0.11) and the forest (1.44 ± 0.26). The observed pattern of functional vulnerability among habitats was slightly higher in the homegarden (75%) than in the forest (71%) and the milpa (69%). Finally, functional over-redundancy was very similar (18-25%) across the three habitats (Table 5). However, given the variability among sites within the same habitat, the functional redundancy, functional over-redundancy, and functional vulnerability did not differ statistically among habitat types (Table 5).

Table 5.

Results of functional analyses based on the functional entities of bees in three habitat types: forest patches (Forest), homegardens and milpas in an agricultural landscape on the Yucatan Peninsula, Mexico. The number of functional entities (FE), functional redundancy (FRed), over-functional redundancy (FOR) and functional vulnerability (FVuln) are shown. Data are mean values ± 1 SE. The results of statistical comparisons between habitats are shown in the last column.

Habitat
Index (Units)	Forest	Homegarden	Milpa	Statistic
FE (Count)	5.25 ± 1.25	6.62 ± 2.13	9 ± 2.39	H = 6.89*
FRed (Species)	1.44 ± 0.26	1.45 ± 0.37	1.58 ± 0.11	H = 1.58
FOR (%)	20 ± 6	18 ± 13	25 ± 2	H = 1.44
FVuln (%)	71 ± 8	75 ± 18	69 ± 3

*p < 0.05.

Similarly, although the functional space of the bee community tends to be larger and with greater functional evenness in the agroecosystems (milpa and homegarden) than in the forest, none of the functional diversity metrics used in this study differed significantly among habitat types (Table 6).

Table 6.

Metrics of functional diversity (functional richness [FRic], functional evenness [FEve], functional divergence [FDiv] and functional dispersion [FDis] of bees in tropical forest patches (Forest) and two traditional agroecosystems (Homegardens and Milpas) in an agricultural landscape of the Yucatan Peninsula, Mexico. Data are mean values ± 1SE. The results of the statistical analyses are shown in the last column, p > 0.05 in all cases.

Habitat
Metric	Forest	Homegarden	Milpa	Statistic
FRic	0.22 ± 0.21	0.31 ± 0.23	0.33 ± 0.23	χ²_2, = 0.71
FEve	0.27 ± 0.12	0.38 ± 0.20	0.36 ± 0.08	χ²₂ = 0.47
FDiv	0.71 ± 0.06	0.68 ± 0.28	0.75 ± 0.08	χ²₂ = 0.77
FDis	0.43 ± 0.28	0.43 ± 0.19	0.50 ± 0.15	χ²₂ = 0.72

Discussion

Contrary to our expectations, we found that traditional agroecosystems were similar to forest remnants in terms of the taxonomic and functional diversity of native bees in an agricultural landscape of the Yucatan Peninsula. The similarity among these landscape units suggests that bees can move across these, use them as habitats, and/or exploit the resources they offer. Moreover, the fact that a proportion of contrasting bee functional groups were found in the main habitats found in the study area (forest, homegarden, and milpa) suggests that management practiced in traditional agroecosystems does not represent an important environmental filter for any bee functional group, even for those bee species that are highly sensitive to anthropogenic disturbance, such as specialists and solitary bees. Given the suitability of traditional agroecosystems as bee habitats and the fact that they represent a dominant cover type in the rural landscape of the Yucatan Peninsula and other regions of the tropics (Altieri & Nicholls, 2017; Villicaña-Hernández et al., 2020; Méndez et al., 2022), we consider that these areas may offer an option of biodiversity-friendly agriculture that could be incorporated into an integrated strategy of landscape management. The maintenance of low-intensive agriculture, along with the conservation of forest remnants, may represent a suitable strategy for the conservation not only of bees but also other components of biodiversity (i.e. birds, mammals) that provide valuable environmental services (Schroth et al., 2004; Gehring et al., 2005).

Taxonomic Diversity

According to the NMDS and PERMANOVA analyses, species composition was similar between the forest patches and the two traditional agroecosystems studied. This is a surprising result because higher diversity was expected in the forest since it represents the original cover of this landscape, harbors higher plant diversity, and is subjected to management of the lowest relative intensity. This finding may be because the management practices used in the agroecosystems did not negatively affect the native bee species and/or the resources they use. For example, herbicides and other agrochemicals were not used in homegardens and manual weeding was the most common practice undertaken to control weeds in the milpas (pers. obs.). On the other hand, although relatively less diverse, the milpa is a polyculture of plants with a very attractive floral display for the bees that occurs synchronously over a small window of time (3 to 4 months). This could generate a magnet effect and attract bees that are not residents of the milpa, but inhabit nearby sites (Holzschuh et al., 2013; Gilpin et al., 2019). For example, Meléndez-Ramírez et al. (2002) reported a richness of 58 bee species associated with some Yucatan cucurbit crops, which are common species in the milpa, while Vides-Borrell et al., (2019) report a richness of 101 bee species in milpa areas of another site on the Yucatan Peninsula (Campeche). Landaverde et al. (2017) suggest that the moderate disturbance that typically exists in and around the Yucatan milpa could also be an element that favors high bee diversity in this agroecosystem due to the presence of a mix of ruderal wild and cultivated species that are attractive to the bees that tend to occur in areas with intermediate disturbance.

The similarity in the taxonomic composition of the bee community between the forest and homegardens is not surprising, as these agroecosystems are similar in terms of plant community structure and diversity (Poot-Pool et al., 2012). It has even been reported that the mixture of native and exotic species in the homegarden, as well as their low-intensity management, may act to favor native fauna, including the bees, as a result of the high availability of floral resources and nesting sites in these agroecosystems (Hass et al., 2018; Schrader et al., 2018; Escobedo-Kenefic et al., 2020; 2022). On the Yucatan Peninsula in particular, homegardens are often located close to patches of forest or regenerating vegetation (Villicaña-Hernández et al., 2020), and such proximity is often positively associated with bee species richness and abundance (Motzke et al., 2016). Sixteen bee species were shared among the three habitat types, suggesting that movement of these species occurs among the habitats. Previous studies have already shown that biodiverse matrices with arboreal elements, such as homegardens, facilitate the movement of vagile organisms, such as bees, between different habitats in heterogeneous landscapes (Kremen & Merenleder, 2018; Von Thaden et al., 2022). For example, Escobedo-Kenefic and colleagues (2022) found no effect of land use unit (forest, semi-natural vegetation, and cropland) on the abundance and species richness of bee flower visitors in a landscape mosaic of Mesoamerica. Thus, agroecosystem and forest patches may function as complementary habitats (Armas-Quiñonez et al., 2020), or as stepping stones, facilitating the movement of bees over open areas (Cadavid-Florez et al., 2020). Although the milpa has no arboreal elements in its interior, some milpas in the study area conserve belts or barriers of trees that extend 5 to 20 m in length around them (known in Mayan as tolché). Added to the large amount of floral resources available, these tree belts could explain the suitability of the milpa as a habitat or foraging site for bees (Briggs et al., 2013; Levy-Tacher et al., 2019; Fonteyne et al., 2023).

Although the general trend shows no differences in bee species composition among habitat types, we found differences in the inverse of Simpson’s diversity index, which is related to a turnover in species dominance. The forest patches were dominated by eusocial bees such as N. perilampoides, which nests in tree trunks (Cab-Baqueiro et al., 2022), and T. fulviventris, which usually nests in rock holes, soil banks, or dead branches (Michener, 2007). The homegardens were dominated by A. metallica, a species that also nests in the ground (Portman et al., 2022), and P. bilineata, which can nest in cultivated tree species such as the Chrysophyllum cainito present in this agroecosystem (Cab-Baqueiro et al., 2022). The milpa was dominated by P. limitaris, an oligolectic species that has been found nesting in soils in the vicinity of crops and is an active visitor of flowers of milpa crops such as species of Cucurbita (Meléndez-Ramírez et al., 2002; Cordeiro et al., 2021).

The number of exclusive species was relatively low in the three habitats under study. The forest harbored some solitary species, such as A. macusolum, which usually nests in trunks and uses plant trichomes to build its nest, and Hylaeus sp., a species reported as rare given its low abundance, which nests in dead stems (Michener, 2007; Gonzalez & Griswold, 2013). Species of Ceratina were found in both the forest and homegardens. These species often nest on dead stems but can also nest in human-made areas (Mikát et al., 2022). In addition, Gaesischia sp., a specialist bee of the Ipomea species that commonly inhabits anthropogenic areas, was found in the homegardens (Araujo et al., 2018). In both the homegardens and milpa, species of the genus Augochlora were recorded. These bees build nests in cavities in soft or rotten wood (Michener, 2007; Dalmazzo & Roig-Alsina, 2012) and their presence in the milpa could therefore only be for foraging and not for nesting. Lasioglossum species were also found in these agroecosystems. These nest in the ground and are favored by open areas (Brown et al., 2020) as well as being important for the pollination of the C. chinense and B. orellana, plants present in these agroecosystems (Landaverde et al., 2017; Rocha et al., 2023). The species exclusive to the milpa were Paratetrapedia sp., a specialist bee that collects floral oils from Malpighiaceae species (Aguiar & Melo, 2011), and leafcutting bees (Megachile sp.), which are often associated with the distribution of Asteraceae and Fabaceae species since they use the leaves for nest construction (Kambli et al., 2017). Both species probably make use of the wild herbaceous resources or wild woody species found in the surroundings of the milpa.

Functional diversity

Functional diversity was similar in the agroecosystems and forest patches. Again, this could be due to the suitability of traditional agroecosystems as a habitat for a functionally diverse community of native bees. In the agricultural landscape of the Yucatan Peninsula, the traditional practices that the Maya have historically carried out have generated a matrix of diverse native crops, surrounded by fragments of highly biodiverse native vegetation (Ford & Topsey, 2019). This produces a mosaic of habitats with agroecosystems that share species (local and cultivated varieties that coexist with their wild relatives) and surrounding vegetation that together provide complementary food resources (Lichtenberg et al., 2017). This allows a diverse set of floral traits that, in turn, attracts a diverse assemblage of bee species, thus favoring the complementarity of plant-pollinator interactions and enhancing crop reproductive success (Badillo-Montaño et al., 2019; Laha et al., 2020; Escobedo-Kenefic et al., 2022; Peralta et al., 2023). In addition to functional similarity, the availability of nesting sites (e.g. trunks, dead stems, herbaceous crops, areas without shade or forest at different successional stages), the null or low use of agrochemicals in these agroecosystems could foster the presence of bee species with the ability to move, but a sensitivity to these substances (Vides-Borrell et al., 2019; Brown et al., 2020; Marcacci et al., 2022).

In contrast to the relative homogeneity across habitats found in the multivariate metrics of functional diversity, we found that the number of functional entities differed among habitats. Again, in contrast to our prediction, we found a greater number of functional entities in the traditional agroecosystems. This is consistent with the evidence that bees are frequently found in areas of abundant floral resources, such as nectar- or pollen-rich crops (Hall et al., 2019) since these habitats offer a wider spectrum of temporarily available resources due to human intervention (Banasza-Cibika & Dylewski, 2021). The bee species found in these habitats share generalist traits, nest in the ground, and are noted for their ability to exploit crop and herbaceous resources; however, there are also species with specialized habits such as squash bees (Peponapis sp.) and some meliponini species that nest in arboreal substrates. These species can thus find resources (mainly food) of importance to their permanence in these traditional agroecosystems.

Limitations and Future Direction

It should be noted that this study was conducted for only part of the year, so it is impossible to guarantee that the pattern described in this study will be maintained during the part of the year not covered by the study. For instance, over an entire year, Meléndez-Ramírez et al. (2016) found two times more bee species than we found in a conserved forest, at a different locality in Yucatan Peninsula. However, although our results may only apply to one particular part of the year, they do show that, when all three habitats are available simultaneously, these offer habitat, food and nesting resources to bees with different mobility and dietary requirements. It is true that milpa is not present throughout the year, however, it has been planted in the study area for centuries and has thus become a seasonal but predictable resource (Ford & Topsey, 2019). Future research exploring seasonal patterns in bee distribution over an entire year may also provide valuable information regarding the seasonal use of habitat by different functional groups of bees.

Although our sampling effort allows a valid comparison among habitats since it was the same in all three habitat types, we may have failed to detect some rare species or those that are perhaps more abundant during the non-sampling period. A longer-term study involving more intensive sampling of study sites (i.e. every month for more than one year) may be required to detect rare species. Our study focused mainly on alpha diversity and future research with an adequate in-depth approach to beta diversity, both taxonomic and functional, could therefore elucidate the complementarity and turnover in the bee community within and among the three main habitats of the studied landscape and other habitats.

Conclusion

In conclusion, taxonomic and functional diversity was similar in the traditional agroecosystems and remnant forest patches, suggesting that, in addition to the forest, traditional agroecosystems can offer important resources to the native bee community and are suitable for contrasting functional groups of native bees.

Implications for Conservation

As demonstrated in this study, homegardens, milpa, and forest all provide food and habitat resources for various functional groups of bees. In addition to their importance in conserving this important group of organisms, traditional agroecosystems contribute to food production and thus ensure food security for the local communities. It is therefore necessary to incorporate the region's traditional agroecosystems into a comprehensive landscape management program. Currently, these agroecosystems face important challenges for their conservation, such as abandonment and the devaluation of agricultural labor, aspects that threaten their existence and the agrobiodiversity they protect. The consideration of agroecosystems as elements that conserve and connect biodiversity and agricultural landscapes is a priority. One way to promote the conservation of this and other biodiversity-rich agroecosystems is to disseminate traditional knowledge and traditional agricultural practices as well as to create local markets for the sale of the products of local agroecosystems (Nimmo et al., 2023). Inclusion of the forest remnants located in the private/communal (i.e. ejido) lands in a scheme of payment for environmental services has also been shown to reduce forest fragmentation in Mexico (Ramírez-Reyes et al., 2018). Agroforestry systems such as home gardens could also be included in a similar scheme of payment already implemented in Mexico as the Sembrando vida program (Gómez-Rodríguez et al., 2023). Preventing the transformation of traditional agroecosystems into more intensive systems, as well as the maintenance of forested areas in the region, should also be a priority.

Supplemental Material

Supplemental Material - Taxonomic and Functional Diversity of Bees in Traditional Agroecosystems and Tropical Forest Patches on the Yucatan Peninsula

Supplemental Material for Taxonomic and Functional Diversity of Bees in Traditional Agroecosystems and Tropical Forest Patches on the Yucatan Peninsula by Laura P. Serralta-Batun, Juan J. Jiménez-Osornio, Virginia Meléndez-Ramírez and Miguel A. Munguía-Rosas in Tropical Conservation Science

Supplemental Material

Supplemental Material - Taxonomic and Functional Diversity of Bees in Traditional Agroecosystems and Tropical Forest Patches on the Yucatan Peninsula

Footnotes

Acknowledgments

E Moo, E Rivero, D Ruiz, A Marote, and G Jiménez helped with the fieldwork. E González-Iturbe produced the map. We thank S Villeger for his advice with the package mFD. We thank the families of Tahdziú and Tzucacab for giving us the access and trust necessary to work in their agroecosystems. K MacMillan translated this manuscript from Spanish to English.

Declaration of Conflicting Interests

The authors declare no potential conflicts of interest concerning the research, authorship, and/or publication of this article.

Funding

The author(s) disclosed receipt of the following financial support for the research, authorship, and/or publication of this article: This study was supported by CONAHCYT (Project: M0037-2019-05-305870), SISTPROY CIRB-FMVZ (2020-0015), and the W.K. Kellogg Foundation (Project: Agencias de desarrollo humano local (ADHL-Alianzas).

ORCID iDs

Laura P. Serralta-Batun

Miguel A.Munguía-Rosas

Supplemental Material

Supplemental material for this article is available online.

References

Aguiar

A. J. C.

Melo

G. A. R.

(2011). Revision and phylogeny of the bee genus Paratetrapedia Moure, with description of a new genus from the Andean Cordillera (Hymenoptera, Apidae, Tapinotaspidini). Zoological Journal of the Linnean Society, 162, 351–442. https://doi.org/10.1111/j.1096-3642.2010.00678.x

Altieri

M. A.

Nicholls

C. I.

(2017). The adaptation and mitigation potential of traditional agriculture in a changing climate. Climatic Change, 140, 33–45. https://doi.org/10.1007/s10584-013-0909-y

Andresen

Arroyo-Rodríguez

Escobar

(2018). Tropical Biodiversity: The Importance of Biotic Interactions for Its Origin, Maintenance, Function, and Conservation. In: Dáttilo

Rico-Gray

(eds) Ecological Networks in the Tropics. Springer, Cham. https://doi.org/10.1007/978-3-319-68228-0_1

Araujo

L. S.

Medina

A. M.

Gimenes

(2018). Pollination efficiency on Ipomoea bahiensis (Convolvulaceae): Morphological and behavioural aspects of floral visitors. Iheringia - Serie Zoologia, 108, e2018012. https://doi.org/10.1590/1678-4766e2018012

Armas-Quiñonez

Ayala-Barajas

Avendaño-Mendoza

Lindig-Cisneros

del-Val

(2020). Bee diversity in secondary forests and coffee plantations in a transition between foothills and highlands in the Guatemalan Pacific Coast. PeerJ, 8, e9257. https://doi.org/10.7717/peerj.9257

Badillo-Montaño

Aguirre

Munguía-Rosas

M. A.

(2019). Pollinator-mediated interactions between cultivated papaya and co-flowering plant species. Ecology and Evolution, 9, 587–597. https://doi.org/10.1002/ece3.4781

Banaszak-Cibicka

Dylewski

Ł.

(2021). Species and functional diversity — A better understanding of the impact of urbanization on bee communities. Science of the Total Environment, 774, 145729. https://doi.org/10.1016/j.scitotenv.2021.145729

Basu

Parui

A. K.

Chatterjee

Dutta

Chakraborty

Roberts

Smith

(2016). Scale dependent drivers of wild bee diversity in tropical heterogeneous agricultural landscapes. Ecology and Evolution, 1, 6983–6992. https://doi.org/10.1002/ece3.2360

Bełcik

Lenda

Amano

Skórka

(2020). Different response of the taxonomic, phylogenetic and functional diversity of birds to forest fragmentation. Scientific Reports, 10, 2045-2322. https://doi.org/10.1038/s41598-020-76917-2

10.

Bentrup

Hopwood

Adamson

N. L.

Vaughan

(2019). Temperate agroforestry systems and insect pollinators: A review. Forests, 1, 14–16. https://doi.org/10.3390/f10110981

11.

Bloom

E. H.

Wood

T. J.

Hung

K. L. J.

Ternest

J. J.

Ingwell

L. L.

Goodell

Kaplan

Szendrei

(2021). Synergism between local- and landscape-level pesticides reduces wild bee floral visitation in pollinator-dependent crops. Journal of Applied Ecology, 58, 1187–1198. https://doi.org/10.1111/1365-2664.13871

12.

Bonfim

F. C. G.

Dodonov

Cazetta

(2021). Landscape composition is the major driver of the taxonomic and functional diversity of tropical frugivorous birds. Landscape Ecology, 36, 2535–2547. https://doi.org/10.1007/s10980-021-01266-y

13.

Briggs

H. M.

Perfecto

Brosi

B. J.

(2013). The role of the agricultural matrix: Coffee management and euglossine bee (Hymenoptera: Apidae: Euglossini) communities in Southern Mexico. Environmental Entomology, 42, 1210–1217. https://doi.org/10.1603/EN13087

14.

Brito

T. F.

Contrera

F. A. L.

Phifer

C. C.

Knowlton

J. L.

Brasil

L. S.

Maués

M. M.

Silva

D. P.

(2018). Effects of habitat type change on taxonomic and functional composition of orchid bees (Apidae: Euglossini) in the Brazilian Amazon. Journal of Insect Conservation, 22, 451–463. https://doi.org/10.1007/s10841-018-0073-9

15.

Brown

Barton

P. S.

Cunningham

S. A.

(2020). Flower visitation and land cover associations of above ground- and below ground-nesting native bees in an agricultural region of south-east Australia. Agriculture, Ecosystems and Environment, 295, 106895. https://doi.org/10.1016/j.agee.2020.106895

16.

Cab-Baqueiro

Ferrera-Cerrato

Quezada-Euán

J. J. G.

Moo-Valle

Vargas-Díaz

A. A.

(2022). Nesting substrates and nest density of stingless bees in the Petenes Biosphere Reserve, Mexico. Acta Biologica Colombiana, 27, 61–69. https://doi.org/10.15446/abc.v27n1.88381

17.

Cadavid-Florez

Laborde

Mclean

D. J.

(2020). Isolated trees and small woody patches greatly contribute to connectivity in highly fragmented tropical landscapes. Landscape and Urban Planning, 196, 103745. https://doi.org/10.1016/j.landurbplan.2020.103745

18.

Camacho-Villa

T. C.

Martinez-Cruz

T. E.

Ramírez-López

Hoil-Tzuc

Terán-Contreras

(2021). Mayan Traditional Knowledge on Weather Forecasting: Who Contributes to Whom in Coping With Climate Change? Frontiers in Sustainable Food Systems, 5, 618453. https://doi.org/10.3389/fsufs.2021.618453

19.

Carrié

Andrieu

Cunningham

S. A.

Lentini

P. E.

Loreau

Ouin

(2017). Relationships among ecological traits of wild bee communities along gradients of habitat amount and fragmentation. Ecography, 40, 85–97. https://doi.org/10.1111/ecog.02632

20.

Cohen

Egerer

Thomas

S. S.

Philpott

S. M.

(2022). Local and landscape features constrain the trait and taxonomic diversity of urban bees. Landscape Ecology, 37, 583–599. https://doi.org/10.1007/s10980-021-01370-z

21.

Cordeiro

G. D.

Liporoni

Caetano

C. A.

Krug

Martínez-Martínez

C. A.

Martins

H. O. J.

Cardoso

R. K. O. A.

Araujo

F. F.

Araújo

P. C. S.

Oliveira

Schlindwein

Warrant

E. J.

Dötterl

Alves-Dos-santos

(2021). Nocturnal bees as crop pollinators. Agronomy, 11, 1014. https://doi.org/10.3390/agronomy11051014

22.

Coutinho

J. G. E.

Hipólito

Santos

R. L. S.

Moreira

E. F.

Boscolo

Viana

B. F.

(2021). Landscape Structure Is a Major Driver of Bee Functional Diversity in Crops. Frontiers in Ecology and Evolution, 9, 624835. https://doi.org/10.3389/fevo.2021.624835

23.

Cruz-Cortés

J. J.

Fraga

J. E.

Munguía-Rosas

M. A.

(2019). Effects of Changes in Traditional Agroecosystems on Vernacular Dwellings: the Occupants’ Perspective. Human Ecology, 47, 553–563. https://doi.org/10.1007/s10745-019-00087-7

24.

Dalmazzo

Roig-Alsina

(2012). Nest structure and notes on the social behavior of Augochlora amphitrite (Schrottky) (Hymenoptera, halictidae). Journal of Hymenoptera Research, 26, 17–29. https://doi.org/10.3897/JHR.26.2440

25.

Díaz

Cabido

(2001). Vive la différence: plant functional diversity matters to ecosystem processes. Trends in Ecology & Evolution, 16, 646-651. https://doi.org/10.1016/S0169-5347(01)02283-2

26.

Escobedo-Kenefic

Landaverde-González

Theodorou

Cardona

Dardón

M. J.

Martínez

Domínguez

C. A.

(2020). Disentangling the effects of local resources, landscape heterogeneity and climatic seasonality on bee diversity and plant-pollinator networks in tropical highlands. Oecologia, 194, 333–344. https://doi.org/10.1007/s00442-020-04715-8

27.

Escobedo-Kenefic

Casiá-Ajché

Q.B.

Cardona

Escobar-González

Mejia-Coroy

Enriquez

Landaverde-González

(2022). Landscape or local? Distinct responses of flower visitor diversity and interaction networks to different land use scales in agricultural tropical highlands. Frontiers in Sustainable Food Systems, 6, 974215. https://doi.org/10.3389/fsufs.2022.974215

28.

Fonteyne

Castillo Caamal

J. B.

Lopez-Ridaura

Van Loon

Espidio Balbuena

Osorio Alcalá

Martínez Hernández

Odjo

Verhulst

(2023). Review of agronomic research on the milpa, the traditional polyculture system of Mesoamerica. Frontiers in Agronomy, 5, 1115490. https://doi.org/10.3389/fagro.2023.1115490

29.

Ford

Topsey

C.E.

(2019). Learning from the ancient maya: conservation of the culture and nature of the maya forest. In: Jameson

J.H.

Musteaţă

(eds) Transforming Heritage Practice in the 21st Century. One World Archaeology. Springer, Cham. https://doi.org/10.1007/978-3-030-14327-5_9

30.

Forrest

J. R. K.

Thorp

R. W.

Kremen

Williams

N. M.

(2015). Contrasting patterns in species and functional-trait diversity of bees in an agricultural landscape. Journal of Applied Ecology, 52, 706–715. https://doi.org/10.1111/1365-2664.12433

31.

Galluzzi

Eyzaguirre

Negri

(2010). Home gardens: Neglected hotspots of agro-biodiversity and cultural diversity. Biodiversity and Conservation, 19, 3635–3654. https://doi.org/10.1007/s10531-010-9919-5

32.

Gehring

Denich

Vlek

P. L. G.

(2005). Resilience of secondary forest regrowth after slash-and-burn agriculture in central Amazonia. Journal of Tropical Ecology, 21, 519–527. https://doi.org/10.1017/S0266467405002543

33.

Gilpin

A. M.

Denham

A. J.

Ayre

D. J.

(2019). Do mass flowering agricultural species affect the pollination of Australian native plants through localized depletion of pollinators or pollinator spillover effects? Agriculture, Ecosystems and Environment, 277, 83–94. https://doi.org/10.1016/j.agee.2019.03.010

34.

Gómez-Rodriguez

G. A.

López-Martínez

J. O.

Špirić

Macario-Mendoza

P. A.

(2023). Local Perceptions in the Implementation of the Sembrando Vida Program in Southern Mexico. Human Ecology, 51, 379–395. https://doi.org/10.1007/s10745-023-00436-7

35.

Gonzalez

V. H.

Griswold

T. L.

(2013). Wool carder bees of the genus Anthidium in the Western Hemisphere (Hymenoptera: Megachilidae): Diversity, host plant associations, phylogeny, and biogeography. Zoological Journal of the Linnean Society, 168, 221–425. https://doi.org/10.1111/zoj.12017

36.

González-Cruz

García-Frapolli

Casas

Dupuy

J. M.

(2015). Responding to disturbances: Lessons from a Mayan social-ecological system. International Journal of the Commons, 9, 831–850. https://doi.org/10.18352/ijc.571

37.

Greenleaf

S. S.

Williams

N. M.

Winfree

Kremen

(2007). Bee foraging ranges and their relationship to body size. Oecologia, 153, 589–596. https://doi.org/10.1007/s00442-007-0752-9

38.

Hall

M. A.

Nimmo

D. G.

Cunningham

S. A.

Walker

Bennett

A. F.

(2019). The response of wild bees to tree cover and rural land use is mediated by species’ traits. Biological Conservation, 231, 1–12. https://doi.org/10.1016/j.biocon.2018.12.032

39.

Hass

A. L.

Liese

Heong

K. L.

Settele

Tscharntke

Westphal

(2018). Plant-pollinator interactions and bee functional diversity are driven by agroforests in rice-dominated landscapes. Agriculture, Ecosystems and Environment, 253, 140–147. https://doi.org/10.1016/j.agee.2017.10.019

40.

Holzschuh

Dormann

C. F.

Tscharntke

Steffan-Dewenter

(2013). Mass-flowering crops enhance wild bee abundance. Oecologia, 172, 477–484. https://doi.org/10.1007/s00442-012-2515-5

41.

Hotspot

Harvey

C. A.

Komar

Chazdon

Ferguson

B. G.

Finegan

Griffith

D. M.

Martínez-Ramos

Morales

Nigh

Soto-Pinto

Van Breugel

Wishnie

(2008). Integrating Agricultural Landscapes with Biodiversity Conservation in the Mesoamerican Hotspot. Biology, 22, 8–15. https://doi.org/10.1111/j.l

42.

Hsieh

T. C.

K. H.

Chao

(2016). iNEXT: an R package for rarefaction and extrapolation of species diversity (Hill numbers). Methods in Ecology and Evolution, 7, 1451–1456. https://doi.org/10.1111/2041-210X.12613

43.

Instituto Nacional de Estadística y Geografía (2020). Número de habitantes. Yucatán. (2020). Consultado el 15 de enero de 2022 a través de. https://cuentame.inegi.org.mx/monografias/informacion/yuc/poblacion/default.aspx

44.

IPBES (2016). The assessment report of the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services on pollinators, pollination and food production. Potts

S. G.

Imperatriz-Fonseca

V. L.

Ngo

H. T.

, (eds). Secretariat of the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services. https://www.ipbes.net/assessment-reports/pollinators

45.

Jost

(2006). Entropy and diversity. Oikos, 113, 363–375. https://doi.org/10.1111/j.2006.0030-1299.14714.x

46.

Kambli

S. S.

Aiswarya

M. S.

Manoj

Varma

Asha

Rajesh

T. P.

Sinu

P. A.

(2017). Leaf foraging sources of leafcutter bees in a tropical environment: implications for conservation. Apidologie, 48, 473–482. https://doi.org/10.1007/s13592-016-0490-2

47.

Kennedy

C. M.

Lonsdorf

Neel

M. C.

Williams

N. M.

Ricketts

T. H.

Winfree

Bommarco

Brittain

Burley

A. L.

Cariveau

Carvalheiro

L. G.

Chacoff

N. P.

Cunningham

S. A.

Danforth

B. N.

Dudenhoffer

J. H.

Elle

Gaines

H. R.

Garibaldi

L. A.

Gratton

Kremen

(2013). A global quantitative synthesis of local and landscape effects on wild bee pollinators in agroecosystems. Ecology Letters, 16, 584–599. https://doi.org/10.1111/ele.12082

48.

Klein

A. M.

Vaissière

B. E.

Cane

J. H.

Steffan-Dewenter

Cunningham

S. A.

Kremen

Tscharntke

(2007). Importance of pollinators in changing landscapes for world crops. Proceedings of the Royal Society B: Biological Sciences, 274, 303-313. https://doi.org/10.1098/rspb.2006.3721

49.

Kohler

Sturm

Sheffield

C. S.

Carlyle

C. N.

Manson

J. S.

(2020). Native bee communities vary across three prairie ecoregions due to land use, climate, sampling method and bee life history traits. Insect Conservation and Diversity, 13, 571–584. https://doi.org/10.1111/icad.12427

50.

Kratschmer

Pachinger

Schwantzer

Paredes

Guzmán

Goméz

J. A.

Entrenas

J. A.

Guernion

Burel

Nicolai

Fertil

Popescu

Macavei

Hoble

Bunea

Kriechbaum

Zaller

J. G.

Winter

(2019). Response of wild bee diversity, abundance, and functional traits to vineyard inter-row management intensity and landscape diversity across Europe. Ecology and Evolution, 9, 4103–4115. https://doi.org/10.1002/ece3.5039

51.

Kremen

Merenlender

A. M.

(2018). Landscapes that work for biodiversity and people. Science, 362, eeau6020. https://doi.org/10.1126/science.aau6020

52.

Laha

Chatterjee

Das

Smith

Basu

(2020). Exploring the importance of floral resources and functional trait compatibility for maintaining bee fauna in tropical agricultural landscapes. Journal of Insect Conservation, 24, 431–443. https://doi.org/10.1007/s10841-020-00225-3

53.

Laliberte

Legendre

(2010). A distance-based framework for measuring functional diversity from multiple traits. Ecology, 91, 299-305. https://doi.org/10.1890/08-2244.1

54.

Landaverde-González

Quezada-Euán

J. J. G.

Theodorou

Murray

T. E.

Husemann

Ayala

Moo-Valle

Vandame

Paxton

R. J.

(2017). Sweat bees on hot chillies: provision of pollination services by native bees in traditional slash-and-burn agriculture in the Yucatán Peninsula of tropical Mexico. Journal of Applied Ecology, 54(6), 1814–1824. https://doi.org/10.1111/1365-2664.12860

55.

Levy-Tacher

S. I.

Ramírez-Marcial

Navarrete-Gutiérrez

D. A.

Rodríguez-Sánchez

P. V.

(2019). Are Mayan community forest reserves effective in fulfilling people’s needs and preserving tree species? Journal of Environmental Management, 245, 16–27. https://doi.org/10.1016/j.jenvman.2019.04.097

56.

Lichtenberg

E. M.

Mendenhall

C. D.

Brosi

(2017). Foraging traits modulate stingless bee community disassembly under forest loss. Journal of Animal Ecology, 86, 1404–1416. https://doi.org/10.1111/1365-2656.12747

57.

Lope-Alzina

D. G.

Howard

P. L.

, (2012). The structure, composition, and functions of homegardens: Focus on the Yucatán Peninsula. Etnoecológica, 9, 17–41.

58.

Magneville

Loiseau

Albouy

Casajus

Claverie

Escalas

Leprieur

Maire

Mouillot

Villéger

(2022). mFD: an R package to compute and illustrate the multiple facets of functional diversity. Ecography, 2021, e05904. https://doi.org/10.1111/ecog.05904

59.

Marcacci

Grass

Rao

V. S.

Kumar S

Tharini

K. B.

Belavadi

V. V.

Nölke

Tscharntke

Westphal

(2022). Functional diversity of farmland bees across rural–urban landscapes in a tropical megacity. Ecological Applications, 32, e2699. https://doi.org/10.1002/eap.2699

60.

Mason

Mouillot

Lee

Wilson

(2005). Functional richness, functional evenness and functional divergence: the primary components of functional diversity. Oikos, 111,112-118. https://doi.org/10.1111/j.0030-1299.2005.13886.x

61.

Meléndez-Ramírez

R. V.

Magaña

R. S.

Parra-Tabla

Ayala

Navarro

(2002). Diversity of native bee visitors of cucurbit crops (Cucurbitaceae) in Yucatán, México. Journal of Insect Conservation, 6, 135 – 147. https://doi.org/10.1023/A:1023219920798

62.

Meléndez-Ramírez

R.V.

Ayala

Delfín G

(2014). Abejas como bioindicadores de perturbaciones en los ecosistemas y el ambiente. In: González Zuarth

Vallarino

Pérez

J. J.

Low

P. A.

(eds.). Bioindicadores: guardianes de nuestro futuro ambiental. Instituto Nacional de Ecología y Cambio Climático (INECC), el Colegio de la Frontera Sur (ECOSUR). https://biblioteca.ecosur.mx/cgi-bin/koha/opac-detail.pl?biblionumber=000054631

63.

Meléndez-Ramírez

Ayala

Delfín González

(2016). Temporal variation in native bee diversity in the tropical sub-deciduous forest of the Yucatan Peninsula, Mexico. Journal Tropical Conservation Science, 9, 718-734. https://doi.org/10.1177/194008291600900210

64.

Méndez

L. M.

Rodríguez

R. K.

Juárez

T. L.

(2022). La experiencia participativa en la elaboración de la estrategia 2020 – 2030 para el Paisaje Forestal Milpero en la Península de Yucatán. In Juárez

T. L.

León

V. J.

Corona

R. N.

(Eds.), Perspectivas del desarrollo territorial inclusivo en el Sur-sureste: desafíos y propuestas transdisciplinarias. Centro de Investigación en Ciencias de Información Geoespacial, A.C. (CentroGeo).

65.

Meneses-Calvillo

L. M.

Meléndez-Ramírez

V. M.

Parra-Tabla

Navarro

(2010). Bee diversity in a fragmented landscape of the Mexican neotropic. Journal of Insect Conservation, 14, 323–334. https://doi.org/10.1007/s10841-010-9262-x

66.

Michener

C.D.

(2007) The bees of the world. The John Hopkins University Press.

67.

Mikát

Fraňková

Benda

Straka

(2022). Evidence of sociality in European small Carpenter bees (Ceratina). Apidologie, 53, 18. https://doi.org/10.1007/s13592-022-00931-8

68.

Millard

Outhwaite

C. L.

Kinnersley

Freeman

Gregory

R. D.

Adedoja

Gavini

Kioko

Kuhlmann

Ollerton

Ren

Z. X.

Newbold

(2021). Global effects of land-use intensity on local pollinator biodiversity. Nature Communications, 12, 2902. https://doi.org/10.1038/s41467-021-23228-3

69.

Motzke

Klein

A. M.

Saleh

Wanger

T. C.

Tscharntke

(2016). Habitat management on multiple spatial scales can enhance bee pollination and crop yield in tropical homegardens. Agriculture, Ecosystems and Environment, 223, 144–151. https://doi.org/10.1016/j.agee.2016.03.001

70.

Mouchet

M. A.

Villéger

Mason

N. W. H.

Mouillot

(2010). Functional diversity measures: An overview of their redundancy and their ability to discriminate community assembly rules. Functional Ecology, 24, 867–876. https://doi.org/10.1111/j.1365-2435.2010.01695.x

71.

Mouillot

Villéger

Parravicini

Kulbicki

Arias-González

J. E.

Bender

Chabanet

Floeter

S. R.

Friedlander

Vigliola

Bellwood

D. R.

(2014). Functional over-redundancy and high functional vulnerability in global fish faunas on tropical reefs. Proceedings of the National Academy of Sciences of the United States of America, 111, 13757–13762. https://doi.org/10.1073/pnas.1317625111

72.

Nigh

Diemont

S. A. W.

(2013). The maya milpa: fire and the legacy of living soil. Frontiers in Ecology and the Environment, 11, e45-e54. https://doi.org/10.1890/120344

73.

Nimmo

E. R.

Nelson

Gómez-Tovar

García

M. M.

Spring

Lacerda

A. E. B.

de Carvalho

A. I.

Blay-Palmer

(2023). Building an Agroecology Knowledge Network for Agrobiodiversity Conservation. Conservation, 3, 491–508. https://doi.org/10.3390/conservation3040032

74.

Odanaka

K. A.

Rehan

S. M.

(2019). Impact indicators: Effects of land use management on functional trait and phylogenetic diversity of wild bees. Agriculture, Ecosystems and Environment, 286. https://doi.org/10.1016/j.agee.2019.106663

75.

Oksanen

Blanchet

F. G.

Kindt

Legendre

Minchin

P. R.

O'Hara

R. B.

Simpson

G. L.

Sólymos

Stevens

M. H. H.

Wagner

(2012). Vegan: Community Ecology Package. Software. https://CRAN.R-project.org/package=vegan

76.

Parys

K. A.

Esquivel

I. L.

Wright

K. W.

Griswold

Brewer

M. J.

(2020). Native pollinators (hymenoptera: Anthophila) in cotton grown in the Gulf South, United States. Agronomy, 10, 698. https://doi.org/10.3390/agronomy10050698

77.

Peralta

Webber

C. J.

Perry

G. L. W.

Stouffer

D. B.

Vázquez

D. P.

Tylianakis

J. M.

(2023). Scale-dependent effects of landscape structure on pollinator traits, species interactions and pollination success. Ecography, 2023, e06453. https://doi.org/10.1111/ecog.06453

78.

Perfecto

Vandermeer

(2008). Biodiversity conservation in tropical agroecosystems: A new conservation paradigm. Annals of the New York Academy of Sciences, 134, 173-200. https://doi.org/10.1196/annals.1439.011

79.

Peters

Keller

Leonhardt

S. D.

(2022). Diets maintained in a changing world: Does land-use intensification alter wild bee communities by selecting for flexible generalists? Ecology and Evolution, 12, e8919. https://doi.org/10.1002/ece3.8919

80.

Pinto

C. E.

Awade

Watanabe

M. T. C.

Brito

R. M.

Costa

W. F.

Maia

U. M.

Imperatriz-Fonseca

V. L.

Giannini

T. C.

(2020). Size and isolation of naturally isolated habitats do not affect plant-bee interactions: A case study of ferruginous outcrops within the eastern Amazon Forest. PLOS ONE, 15: e0238685. https://doi.org/10.1371/journal.pone.0238685

81.

Poot-Pool

Van Der Wal

Flores-Guido

Manuel Pat-Fernández

Esparza-Olguín

(2012). Economic stratification differentiates home gardens in the Maya Village of Pomuch, Mexico. Economic Botany, 66, 264-275. https://doi.org/10.1007/s12231-012-9206-3

82.

Portman

Z. M.

Arduser

Lane

I. G.

Cariveau

D. P.

(2022). A review of the Augochloropsis (Hymenoptera, Halictidae) and keys to the shiny green Halictinae of the midwestern United States. ZooKeys, 1130, 103–152. https://doi.org/10.3897/zookeys.1130.86413

83.

Prado

S. G.

Ngo

H. T.

Florez

J. A.

Collazo

J. A.

(2017). Sampling bees in tropical forests and agroecosystems: a review. Journal of Insect Conservation, 21,753–770. https://doi.org/10.1007/s10841-017-0018-8

84.

R Core Team (2022). R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. URL https://www.R-project.org/

85.

Rader

Bartomeus

Tylianakis

J. M.

Laliberté

(2014). The winners and losers of land use intensification: Pollinator community disassembly is non-random and alters functional diversity. Diversity and Distributions, 20, 908–917. https://doi.org/10.1111/ddi.12221

86.

Rahimi

Barghjelveh

Dong

(2022). Amount, distance-dependent and structural effects of forest patches on bees in agricultural landscapes. Agriculture and Food Security, 11, 10. https://doi.org/10.1186/s40066-022-00360-x

87.

Ramirez-Reyes

Sims

Potapov

Radeloff

V. C.

(2018). Payments for ecosystems services in Mexico reduce forest fragmentation. Ecological Applications, 28, 1982-1997. https://doi.org/10.1002/eap.1753

88.

Rajagopal

Cuevas Sánchez

J. A.

del Moral

J. B.

Montejo

D. A.

Hernández

T. G.

Lozano

J. L. R.

(2021). The scope and constraints of homegardens for sustainable development: A review. Tropical and Subtropical Agroecosystems, 24, 1-24. https://doi.org/10.56369/tsaes.3487

89.

Rico-Gray

García-Franco

J.G.

(1992). Vegetation and soil seed bank of successional stages in tropical lowland deciduous forest. Journal of Vegetation Science, 13, 617-624. https://doi.org/10.2307/3235828

90.

Rocha

F. H.

Peraza

D. N.

Medina

Quezada-Euán

J. J. G.

(2023). Pollination service provided by honey bees to buzz-pollinated crops in the Neotropics. PLOS ONE, 18, e0280875. https://doi.org/10.1371/journal.pone.0280875

91.

Rodríguez S

Pérez-Giraldo

L. C.

Vergara

P. M.

Carvajal

M. A.

Alaniz

A. J.

(2021). Native bees in Mediterranean semi-arid agroecosystems: Unravelling the effects of biophysical habitat, floral resource, and honeybees. Agriculture, Ecosystems and Environment, 307, 107188. https://doi.org/10.1016/j.agee.2020.107188

92.

Rodríguez-Robayo

K. J.

Méndez-López

M. E.

Molina-Villegas

Juárez

(2020). What do we talk about when we talk about milpa? A conceptual approach to the significance, topics of research and impact of the mayan milpa system. Journal of Rural Studies, 77, 47–54. https://doi.org/10.1016/j.jrurstud.2020.04.029

93.

Schrader

Franzén

Sattler

Ferderer

Westphal

(2018). Woody habitats promote pollinators and complexity of plant–pollinator interactions in homegardens located in rice terraces of the Philippine Cordilleras. Paddy and Water Environment, 16, 253–263. https://doi.org/10.1007/s10333-017-0612-0

94.

Schroth

B da Fonseca

G. A.

Harvey

C. A.

Gascon

Vasconcelos

H. L.

Izac

A.-M. N.

(2004). Agroforestry and Biodiversity Conservation in Tropical Landscapes. Washington USA: Island Press.

95.

Serralta-Batun

Jiménez-Osornio

Munguía-Rosas

Rodríguez-Robayo

(2024). Amenazas al paisaje agrícola tradicional del sur de Yucatán, México. Una mirada desde el análisis socioecológico. Revista de Economia e Sociologia Rural, 62(1), e265073. https://doi.org/10.1590/1806-9479.2022.265073

96.

Siviter

Bailes

E. J.

Martin

C. D.

Oliver

T. R.

Koricheva

Leadbeater

Brown

M. J. F.

(2021). Agrochemicals interact synergistically to increase bee mortality. Nature, 596, 389–392. https://doi.org/10.1038/s41586-021-03787-7

97.

Stavert

J. R.

Liñán-Cembrano

Beggs

J. R.

Howlett

B. G.

Pattemore

D. E.

Bartomeus

(2016). Hairiness: The missing link between pollinators and pollination. PeerJ, 12, e2779. https://doi.org/10.7717/peerj.2779

98.

Török

Gallé

Batáry

(2022). Fragmentation of forest-steppe predicts functional community composition of wild bee and wasp communities. Global Ecology and Conservation, 33, e01988. https://doi.org/10.1016/j.gecco.2021.e01988

99.

Twerd

Sobieraj-Betlińska

(2020). Wild bee (Apiformes) communities in contrasting habitats within agricultural and wooded landscapes: implications for conservation management. Agricultural and Forest Entomology, 22, 358–372. https://doi.org/10.1111/afe.12391

100.

Vides-Borrell

Porter-Bolland

Ferguson

B. G.

Gasselin

Vaca

Valle-Mora

Vandame

(2019). Polycultures, pastures and monocultures: Effects of land use intensity on wild bee diversity in tropical landscapes of southeastern Mexico. Biological Conservation, 236, 269–280. https://doi.org/10.1016/j.biocon.2019.04.025

101.

Villéger

Mason

N. W. H.

Mouillot

(2008). New multidimensional functional diversity indices for a multifaceted framework in functional ecology. Ecology, 89(8), 2. https://doi.org/10.1890/07-1206.1

102.

Villicaña-Hernández

G. J.

Martínez-Natarén

D. A.

Álvarez-Espino

R. X.

Munguía-Rosas

M. A.

(2020). Seed Rain in a Tropical Dry Forest and Adjacent Home Gardens in the Yucatan. Tropical Conservation Science, 13, 1-9. https://doi.org/10.1177/1940082920974599

103.

Vogel

Poveda

Iverson

Boetzl

F. A.

Mkandawire

Chunga

T. L.

Küstner

Keller

Bezner Kerr

Steffan‐Dewenter

(2023). The effects of crop type, landscape composition and agroecological practices on biodiversity and ecosystem services in tropical smallholder farms. Journal of Applied Ecology, 160, 859-874. https://doi.org/10.1111/1365-2664.14380

104.

Von Thaden

Salazar-Arteaga

Laborde

Estrada-Contreras

Romero-Uribe

(2022). Arboreal elements of the agricultural matrix as structural connecting devices in fragmented landscapes – A case study in the Los Tuxtlas Biosphere Reserve. Ecological Engineering, 179, 106633. https://doi.org/10.1016/j.ecoleng.2022.106633

105.

Williams

N. M.

Crone

E. E.

Roulston

T. H.

Minckley

R. L.

Packer

Potts

S. G.

(2010). Ecological and life-history traits predict bee species responses to environmental disturbances. Biological Conservation, 143, 2280–2291. https://doi.org/10.1016/j.biocon.2010.03.024

106.

Zamora-Crescencio

García-Gil

Flores-Guido

Ortiz

J.J.

(2008). Estructura y composición florística de la selva mediana subcaducifolia en el sur del estado de Yucatán, México. Polibotánica, 26, 39-66.

107.

Zhang

Zang

(2021). Tropical forests are vulnerable in terms of functional redundancy. Biological Conservation, 262, 3. https://doi.org/10.1016/j.biocon.2021.109326

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