Ant Communities Resist Even in Small and Isolated Gypsum Habitat Remnants in a Mediterranean Agroecosystem

Abstract

Natural and seminatural habitat remnants play a crucial ecological role in intensified agroecosystems. Assumptions on the conservation value of small and poorly connected fragments in a hostile matrix come from generalization obtained from a limited number of taxa, mostly plants, and vertebrates. To date, few studies have analyzed the effect of fragmentation on ant communities in Mediterranean agroecosystems, despite the importance of this group of animals on several key ecosystem functions and services. Here, we analyze the effects of fragment area and connectivity on ant communities in gypsum outcrops in a large cereal agroecosystem of Central Spain. Ant communities were described by their species composition, abundance (total number of occurrences), and number of species, standardized both by area (species density), and abundance (species richness). Observed number of species was relatively high in comparison with other studies in the Mediterranean, and we found no effects of fragment characteristics on species density, species richness and species composition, which implies that even small and isolated patches do have a value for ant conservation. Moreover, total number of occurrences were higher for smaller and more isolated fragments. This finding contrasts with the results reported for other taxa in similar gypsum habitats and suggests that certain ant traits and strategies make them particularly resistant to fragmentation and capable to take advantage of small habitat patches. Given the important ecological role played by ants, we recommend the preservation of these small habitat fragments in the management plans of agroecosystems in these drylands, especially in those cases in which intensification of agricultural practices greatly diminish natural habitat availability.

Full text

ORIGINAL RESEARCH
published: 16 March 2021
doi: 10.3389/fevo.2021.619215
Frontiers in Ecology and Evolution | www.frontiersin.org 1March 2021 | Volume 9 | Article 619215
Edited by:
Anna Sofie Persson,
Lund University, Sweden
Reviewed by:
Inge Armbrecht,
University of Valle, Colombia
Barbara Baraibar,
Universitat de Lleida, Spain
*Correspondence:
Francisco M. Azcárate
fm.azcarate@uam.es
Specialty section:
This article was submitted to
Biogeography and Macroecology,
a section of the journal
Frontiers in Ecology and Evolution
Received: 19 October 2020
Accepted: 14 January 2021
Published: 16 March 2021
Citation:
Azcárate FM, Alameda-Martín A,
Escudero A and Sánchez AM (2021)
Ant Communities Resist Even in Small
and Isolated Gypsum Habitat
Remnants in a Mediterranean
Agroecosystem.
Front. Ecol. Evol. 9:619215.
doi: 10.3389/fevo.2021.619215
Ant Communities Resist Even in
Small and Isolated Gypsum Habitat
Remnants in a Mediterranean
Agroecosystem
Francisco M. Azcárate 1,2
*, Aitor Alameda-Martín 3, Adrián Escudero 4and Ana M. Sánchez 4
1Terrestrial Ecology Group, Department of Ecology, Universidad Autónoma de Madrid, Madrid, Spain, 2Centro de
Investigación en Biodiversidad y Cambio Global (CIBC-UAM), Universidad Autónoma de Madrid, Madrid, Spain,
3Departamento de Edafología y Química Agrícola, Universidad de Almería, Almería, Spain, 4Área de Biodiversidad y
Conservación, Universidad Rey Juan Carlos, Móstoles, Spain
Natural and seminatural habitat remnants play a crucial ecological role in intensified
agroecosystems. Assumptions on the conservation value of small and poorly connected
fragments in a hostile matrix come from generalization obtained from a limited number
of taxa, mostly plants, and vertebrates. To date, few studies have analyzed the effect
of fragmentation on ant communities in Mediterranean agroecosystems, despite the
importance of this group of animals on several key ecosystem functions and services.
Here, we analyze the effects of fragment area and connectivity on ant communities in
gypsum outcrops in a large cereal agroecosystem of Central Spain. Ant communities
were described by their species composition, abundance (total number of occurrences),
and number of species, standardized both by area (species density), and abundance
(species richness). Observed number of species was relatively high in comparison with
other studies in the Mediterranean, and we found no effects of fragment characteristics
on species density, species richness and species composition, which implies that
even small and isolated patches do have a value for ant conservation. Moreover, total
number of occurrences were higher for smaller and more isolated fragments. This finding
contrasts with the results reported for other taxa in similar gypsum habitats and suggests
that certain ant traits and strategies make them particularly resistant to fragmentation and
capable to take advantage of small habitat patches. Given the important ecological role
played by ants, we recommend the preservation of these small habitat fragments in the
management plans of agroecosystems in these drylands, especially in those cases in
which intensification of agricultural practices greatly diminish natural habitat availability.
Keywords: agroecosystems, ants, biodiversity conservation, drylands, fragmentation, gypsum habitats
INTRODUCTION
There is consensus on the necessity of maintaining a certain amount of remnants of natural or
semi-natural habitats in extensive agroecosystems, which act as reservoirs of biodiversity and
provide key ecosystem services (Altieri, 1999; Tscharntke et al., 2008; Kennedy et al., 2013). It is
not clear, however, to what extent small, isolated fragments can fulfill these functions. According to

Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
the most classical view of fragmentation, there is a threshold
in which remnants are too small and isolated from each other,
inducing population declines, increasing extinction risk, and
exacerbating species loss (Andrén, 1994; Ewers and Didham,
2006). Fragmentation is therefore considered by many authors
one of the most pervasive drivers of biodiversity loss (Fletcher
et al., 2018). Some of the mechanisms proposed to explain
these negative processes are related to the higher vulnerability
to environmental and demographic stochasticity for small and
poorly connected populations, or the reduction in colonization
and rescue events, under a metapopulation perspective, which
can become too infrequent to overcome local extinctions
(Andrén, 1994; Ewers and Didham, 2006). These negative effects
would emerge as a function of the specific characteristics of
the landscapes, the habitat configuration and the biological
groups involved (Villard and Metzger, 2014). Thus, for example,
megafauna requires large patches to maintain viable populations,
while in the case of invertebrates very small patches could be
enough to host viable populations (Ewers and Didham, 2006).
Recently, however, it has been questioned whether
fragmentation necessarily means a real loss of species. Fahrig
(2003, 2017) considers that we should only attribute to the
so-called fragmentation per se, those effects on populations and
communities independent of the effects of habitat amount and,
definitively, resource availability. The controversy is significant
for conservation, since the lower conservation value assigned to
small patches have derived in the protection of few large rather
than many small fragments (“single large vs. several small,” or
SLOSS debate; Fahrig et al., 2019). Much of the controversy has
to do with the difficulties of distinguishing between the effects
expected by simple species—area relationships or sampling
effects, and those produced by mechanisms of another nature
(Gotelli and Colwell, 2001; Whittaker et al., 2001). To face these
confounding effects, Whittaker et al. (2001) propose to fix by
design the sampling area, and recommend the use of species
density to make comparisons, while (Gotelli and Colwell, 2001,
2011) propose the use of rarefaction curves to standardize
datasets to a common number of samples, individuals or
occurrences to assess species richness. Both species density and
species richness are worth to be explicitly considered in this
context, since while species density is of interest for conservation
purposes, species richness is appropriated for testing models
and theoretical predictions, and along with species abundance,
is critical for determining species density (Gotelli and Colwell,
2001).
Agriculture intensification and maintenance of large and
homogeneous agricultural fields has induced a critical and
massive destruction of habitat and biodiversity loss (Emmerson
et al., 2016). In this context, the value in conservation terms, but
also as providers of ecosystem services of habitat remnants, is
critical (Haddad et al., 2015). Although it exists a wide consensus
that agroecosystems should maintain and extend the number
and extension of natural and seminatural habitats fragments for
maximizing the services provided, there is no information of the
role these fragments have for many biological groups.
Despite their enormous biomass and importance in the
functioning of terrestrial ecosystems (Folgarait, 1998; Del Toro
et al., 2012), ants as a group have been rarely studied in relation
to fragmentation. This is significant since ground dwelling ants,
as invertebrates living in sessile colonies, probably respond to
fine-scale effects better than larger or more mobile organisms.
To our knowledge, however, there are no published studies
specifically aimed to evaluate the effect of fragmentation on
ants in Mediterranean agroecosystems, despite the ecological
importance of ants in Mediterranean regions (Baraibar et al.,
2009; Silvestre et al., 2019) where they provide important
ecosystem services (Comas et al., 2016; Baraibar et al., 2017).
Most of the existing studies on fragmentation and ants have
been conducted in tropical regions (Armbrecht et al., 2001; Assis
et al., 2018; Santos et al., 2018), forests (Melliger et al., 2018), or
grasslands (Golden and Crist, 2000; Brascheler and Baur, 2003;
Dauber et al., 2006). Few generalizations can be obtained from
them, since apart from their variability in the target ecosystem,
they also differ greatly in the spatial scale, ranging from <1 m
(Brascheler and Baur, 2003) to several thousand square meters
(Melliger et al., 2018).
Here, we aim to evaluate the effects of patch size and
connectivity on ant biodiversity in an extensive cereal
agroecosystem in the drylands of Central Spain. We assume that
larger patches have a higher total number of species because
of a simple random sampling effect, therefore this variable will
not be measured in this study. Instead, we fix sampling area in
each patch, and focus on species density, species richness and
total abundance. If small or isolated patches add negative effects
for ant populations other than simple random sampling effects,
then we expect to find lower values of species density, richness
and abundance. Conversely, if low and isolated patches are
not affected by negative affects additional to random sampling
effects, ant species richness, abundance and density should be
similar to larger and better connected patches.
Our approach will provide additional conservation criteria for
these valuable ecosystems, contributing to preserve their diversity
and therefore their functionality and ecosystem services.
MATERIALS AND METHODS
Study Area
The study area is in central Spain, near the locality of Belinchón
(Cuenca, Spain), at an altitude of ∼700 a.s.l (Figure 1). The
climate is continental Mediterranean, with an average annual
temperature of ca 14◦C, rainfall of ca 500 mm per year,
severe summer drought and cold winters. The lithology of
the area is mostly composed of tertiary evaporites, and the
relief is gentle, with alternating hills and plains. Soils are
gypseous (i.e. soils which main component is gypsum; Herrero
and Porta, 2000), which imposes harsh conditions for the
development of vegetation (Escudero et al., 2015). In spite of
this, and especially from the mid-20th century (Matesanz et al.,
2009), the expansion of rain-fed agriculture has progressively
fragmented the landscape, which now consists of a mosaic
of natural vegetation remnants interspersed in a matrix of
dry herbaceous croplands, mostly cereal (Figure 1). Natural
vegetation in these remnants is composed by dwarf shrublands
dominated by creeping and cushion-like specialized gypsophytes
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Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
FIGURE 1 | (A) Map showing the distribution of gypsum outcrops of the Iberian Peninsula, re-drawn from Escavy et al. (2012), and the study area location (black dot).
(B) Studied habitat remnants. The black dots indicate the centroid of each remnant, and the yellow star indicates the village of Belinchón. Note that the area includes
other habitat remnants that were not sampled. However, all remnants were considered for the calculation of the connectivity index of the sampled ones. (C) General
view of the study area showing a habitat remnant on the forefront with typical gypsum shrubby vegetation, another habitat remnant in the background, and a fallow
field in the middle.
(e.g., Helianthemum squamatum (L.) Dum.-Cours., Lepidium
subulatum L., Centaurea hyssopifolia Vahl., Gypsophila struthium
L) and the tussock-forming grass Stipa tenacissima L. Perennials
cover is around 30% and alternates with clearings partially
covered by a biological soil crust, dominated by crustose lichens
and fairly diverse annual plant communities (Luzuriaga et al.,
2018).
Data Collection
Twenty natural habitat remnants were selected, separated from
each other by a continuous cropland matrix (Figure 1,Table 1)
and including a representative sample of existent range of size
and connectivity in the study area. Fragment characteristics (area,
connectivity) were measured on orthophotos from 2018, using
the software QGIS 3.2.1. Bonn (QGIS.org, 2018). Fragment area
ranged from 1,065 to 290,828 m2, and was log-transformed for
all the analysis, given the strong skewness of its distribution. Log
(area) ranged from 3.03 to 5.46 (mean =4.07, s.d. =0.72).
Connectivity was estimated using the index C proposed
by Tremlová and Münzbergová (2007), which accounts for
the number of surrounding fragments, considering their area
weighted by their distance to the target fragment. Because of the
highly dispersed values obtained, we log10 transformed the index,
following a modification by Matesanz et al. (2015). The modified
index can adopt negative values for lowly connected remnants,
and is calculated as follows:
Cj=log10
n
X
k=1
Pk/d2
jk,j6= k
where Cjis fragment jconnectivity, kthe fragments surrounding
fragment j(in our case, within a radius of 1 km from the center of
fragment j), Pkis fragment karea (in square meters), and djk is the
distance between the centroids of fragments jand k(in meters).
Connectivity index ranged from −0.42 to 0.80 (mean =0.41, s.d.
=0.35).
Ant sampling was carried out in July 2018. In each habitat
remnant a 10 ×15 m plot was located in a homogeneous
vegetation area at a distance of 10 m from the cultivated area, so
fixing the possible edge effect (Golden and Crist, 2000; Brascheler
and Baur, 2003). On each plot, we placed 12 pitfall traps of 2.5 cm
in diameter and 5 cm deep, forming a grid of 3 ×4 rows, with a
separation of 5 m between traps. The traps contained a solution
composed of 70% ethanol and 30% ethylene glycol (Azcárate and
Peco, 2012; Silvestre et al., 2019). The traps were kept in the field
for 7 days, and after their collection, all ant workers were sorted
in the laboratory, and identified to the species level.
Data Analysis
We characterized each plot by its species density, species richness
and overall abundance, following Gotelli and Colwell (2011).
Species density per plot was assimilated to the total number
of species captured in the set of 12 pitfall traps. Total number
of occurrences was used as an indicator of overall abundance
(Longino et al., 2002). Each species present in one trap was
counted as one occurrence, regardless the number of workers
captured in the trap, and total number of occurrences was
calculated as the sum of the occurrences of all the species present
in one plot. We preferred the total number of occurrences instead
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Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
TABLE 1 | Fragment characteristics and ant community metrics obtained for the
20 sampling units considered in this study.
Area (m2) Connectivity W O S S44 ICE S/ICE
F1 2466.70 −0.11 476 92 14 13.24 14 99.9%
F2 6700.50 −0.42 642 96 13 12.01 13.47 96.5%
F3 18403.37 −0.25 221 59 13 11.61 23.99 54.2%
F4 2682.67 0.40 600 60 11 10.65 11.54 95.3%
F5 2395.05 0.77 340 52 9 8.75 12.54 71.8%
F6 9299.88 0.79 329 51 10 9.92 10.33 96.8%
F7 279539.94 −0.02 391 55 10 9.82 10.45 95.7%
F8 3097.74 0.38 316 69 11 10.31 12.12 90.8%
F9 1065.62 0.80 254 70 12 11.46 12.53 95.8%
F10 38232.22 0.70 249 70 13 11.91 14.74 88.2%
F11 14106.15 0.71 208 53 11 10.33 16.25 67.7%
F12 5619.14 0.74 228 75 13 11.67 15.25 85.2%
F13 177775.24 0.47 279 65 11 10.7 11.3 97.3%
F14 24792.99 0.45 261 63 12 11.15 14.53 82.6%
F15 11841.91 0.39 258 63 11 10.45 12.15 90.5%
F16 9318.67 0.58 289 66 14 12.47 17.43 80.3%
F17 290828.16 0.35 198 44 9 9 10.18 88.4%
F18 41693.84 0.38 283 65 17 15.15 25.81 65.9%
F19 6920.26 0.48 243 53 10 9.64 11.40 87.7%
F20 1469.28 0.52 464 81 12 11.3 12.53 95.8%
Total 6,529 23
Area considers the whole remnant surface. Connectivity was estimated using the C index
(Tremlová and Münzbergová, 2007) modified by Matesanz et al. (2015). Ant community
metrics were measured from 12 pit-fall traps arranged regularly in a 10 ×15 m plot within
the remnant. W, Total number of workers; O, Total number of occurrences; S, Species
density; S44, Species richness estimated for 44 occurrences; ICE, Incidence-based
coverage estimator.
of the total number of individuals, because in ants the number
of workers captured by pitfall traps is extremely influenced by
slight differences in the distance to foraging trails or nests, which
results in a strong spatial clumping of individuals within traps.
Anyhow, some analyses were replicated using total number of
individuals, obtaining similar results to those reported in this
study (Supplementary Table 1). Species richness was estimated
as the expected number of species for a given number of
randomly sampled occurrences. To assess it, we built occurrence-
based rarefaction curves (Gotelli and Colwell, 2011), averaging
500 resamples with replacement, and then estimated for each
plot the mean species richness expected for the lowest number
of occurrences observed in any of the 20 plots. As a complement
to these variables, and to check the capacity of our sampling
to cover a representative proportion of the assemblage richness,
we calculated the incidence-based coverage estimator (ICE), an
asymptotic estimator of species richness based on incidence
data (Gotelli and Colwell, 2011). Rarefaction analysis and ICE
estimation were done using the software EstimateS v. 9.1.0
(Colwell, 2013).
We built generalized linear models for species density, total
number of occurrences and species richness, as function of log
(area) and connectivity, including their interaction. We used
TABLE 2 | Ant species identified in this study, with information on their subfamily,
and general data on the frequency & abundance observed in this study.
Subfamily Plots Traps N◦
(0–20) (0–240) workers
Aphaenogaster iberica Emery. 1908 M 7 36 90
Aphaenogaster senilis Mayr. 1853 M 14 123 540
Camponotus foreli Emery. 1881 F 19 114 246
Cataglyphis iberica (Emery. 1906) F 20 206 955
Crematogaster auberti Emery. 1869 M 18 142 1,117
Goniomma blanci (André. 1881) M 5 5 6
Goniomma hispanicum (André. 1883) M 1 3 6
Messor barbarus (Linnaeus. 1767) M 18 93 429
Messor bouvieri Bondroit. 1918 M 16 69 212
Messor capitatus (Latreille. 1798) M 4 10 18
Oxyopomyrmex saulcyi Emery. 1889 M 4 6 7
Pheidole pallidula (Nylander. 1849) M 9 56 1,024
Plagiolepis pygmaea (Latreille. 1798) F 8 12 24
Plagiolepis schmitzii Forel. 1895 F 15 39 116
Proformica sp1 F 19 123 302
Solenopsis sp1 M 1 1 1
Tapinoma erraticum (Latreille. 1798) D 1 3 20
Tapinoma gr. nigerrimum Nylander. 1856 D 1 11 87
Temnothorax formosus (Santschi. 1909) M 13 43 96
Temnothorax universitatis (Espadaler. 1997) M 3 4 10
Tetramorium gr. caespitum (Linnaeus. 1758) M 18 122 697
Tetramorium forte Forel. 1904 M 11 58 399
Tetramorium semilaeve André. 1883 M 11 23 127
Subfamilies: D, Dolichoderinae; F, Formicinae; M, Myrmicinae. Plots, total number of plots
in which the species occurred; Traps, total number of traps in which the species occurred,
considering the 20 plots; N◦of workers, total number of workers collected in the set of
20 plots.
Poisson errors and log link functions for species density and total
number of occurrences, and Gaussian errors and identity link
function for species richness. This latter variable is non-integer,
due to the fact that it is estimated from the rarefaction curves.
We identified the best descriptive models using the Akaike
information criterion corrected for small samples (AICc). Models
were estimated with package MuMIn (Barto´
n, 2009) of R 4.0.0.
We performed a PERMANOVA analysis in order to check
whether there were differences in species composition related to
area, connectivity and/or their interaction. We used frequency
data to estimate species composition in each habitat fragment and
calculated dissimilarity matrices based on Bray Curtis distance.
Significance of the model was tested using a Monte-Carlo test
with 9,999 permutations. Finally, we performed non-metric
multidimensional scaling (NMDS) to visualize the differences in
the species compositions. PERMANOVA and NMDS analyses
were performed with the vegan package in R 4.0.0 (Oksanen et al.,
2015).
RESULTS
We collected 6,529 ant workers belonging to 23 species, all of
them native (Table 2). The recorded number of species per plot
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Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
TABLE 3 | Parameter estimates and P-values for the GLM (Poisson distribution,
log link function) built for species density as a function of Log (Area), Connectivity
and the interaction Log (Area) x Connectivity.
Estimate Std. error z-value Pr(>|z|)
Intercept 3.025 0.651 4.646 <0.0001
Log (Area) −0.126 0.158 −0.797 0.425
Connectivity −1.110 1.387 −0.800 0.424
Log (Area) ×Connectivity, 0.252 0.351 0.720 0.472
N=20 fragments.
TABLE 4 | Akaike Information Criterion corrected for small sample sizes (AICc)
assessed for the models estimated for species density, total number of
occurrences and species richness (for 44 occurrences) as function of the different
combinations of the predictors Log (Area) Connectivity and Log (Area) x
Connectivity.
Predictors Species density Total number Species richness
occurrences
Log (Area), Connectivity,
Log (Area) ×Connectivity
101.69 154.28 81.88
Log (Area), Connectivity 99.06 155.44 80.53
Log (Area) 96.66 164.98 78.95
Connectivity 96.41 166.33 77.91
Null model (no predictors) 76.30 172.24 76.39
In bold the AICc indicating the most plausible models estimated for the different dependent
variables. N =20 fragments. Models were GLM with Poisson distribution and log link
function for Species density and Total number of occurrences, and Gaussian distribution
and identity function for Species richness.
TABLE 5 | Parameter estimates and P-values for the GLM (Poisson distribution,
log link function) built for total number of occurrences as a function of Log (Area),
Connectivity and the interaction Log (Area) x Connectivity.
Estimate Std. error z-value Pr(>|z|)
Intercept 5.339 0.278 19.167 <0.0001
Log (Area) −0.262 0.069 −3.821 0.0001
Connectivity −1.492 0.595 −2.507 0.0122
Log (Area) ×Connectivity 0.312 0.152 2.057 0.0397
(species density) was, on average, 86% of the species richness
expected by the ICE estimator (Table 1). The ratio between the
number of detected and expected species showed no correlation
with connectivity (Pearson r=0.01) or with the area of the
remnant (Pearson r=0.17), so we can assume no sampling bias
associated with the fragmentation condition of the remnants.
Species density ranged from 9 to 17 species per plot (average
11.80, s.d. =1.94), and did not respond to any of the predictors
(Table 3). None of the combinations of the predictors produced
a plausible model, according to the AICc (Table 4).
The total number of occurrences per plot ranged from 44 to 96
(mean =65.10, s.d. =13.33), and responded significantly to log
(Area) and connectivity both in in the complete model (Table 5)
and in the model without interaction. AICc was very similar for
the two models (Table 4). The percentage of explained residual
deviance (D2) was 52.7% for the complete model and 44.1%
for the model without interaction. Opposite to our expectations,
the effects of both connectivity and Log (Area) were negative
(Figure 2), suggesting that neither the reduction of fragment
size nor the increase in isolation imply a loss of habitat quality
for ants. Moreover, the significant interaction suggested that
the increase in total number of occurrences observed for small
remnants was greater if they were more isolated.
Species richness was estimated for a sample size rarefied down
to 44 occurrences (the lowest observed value for any of the plots),
using individual-based rarefaction curves (Figure 3). Estimated
values ranged between 8.75 and 15.15 species (mean =11.08, s.d.
=1.49). Similarly to species density, rarefied species richness did
not respond to any of the considered predictors (Table 6), and the
null model was the one that showed the smallest AICc (Table 4)
The NMDS/PERMANOVA approaches also did not reveal
any trend in relation to patch size or connectivity (Table 7,
Supplementary Figure 1).
DISCUSSION
According to the expectations of the classical theory of ecological
fragmentation, too small and isolated habitat remnants should
undergo an accelerated species loss (Andrén, 1994; Ewers and
Didham, 2006). In our case study, however, neither ant species
density, nor richness or community composition, responded to
fragment size or connectivity. A possible first explanation could
be related to the fact that processes causing species loss in habitat
remnants (Levins, 1969; Andrén, 1994;Ewers and Didham, 2006)
are not necessarily linear, and, even more, they could appear
below certain thresholds of size or isolation or combination
(Villard and Metzger, 2014). Therefore, the hypothetical negative
effects of fragmentation on ant communities would only be
detectable in remnants smaller and more isolated than those here
considered. Furthermore, the total number of occurrences not
only did not decrease but tended to increase in small and less
connected patches, making clear that these fragments do not
undergo any loss of quality as ant habitats.
These results contrast with those obtained for other organisms
in the same region and type of agroecosystem. For example,
Matesanz et al. (2009) and Matesanz et al. (2015) observed
clear population-level effects for plants, and at the community
level, other studies have reported reductions in species density
for plants (Luzuriaga et al., 2018) and lichens (Concostrina-
Zubiri et al., 2018). Fragmentation in gypsum agroecosystems
also seems to negatively affect species richness of pollinator
communities, as well as the frequency, diversity and topology of
pollination interactions and network metrics (Santamaría et al.,
2018). Fragmentation in this type of agroecosystem in drylands
also influenced the risk of seed predation, although its effect was
not straightforward and it appeared as highly dependent on each
plant-animal interaction (Rabasa et al., 2005, 2009; Moncalvillo
et al., 2021). Although we do not have studies specifically aimed at
analyzing the effects of fragmentation on ants in Mediterranean
agroecosystems, it has been observed that the presence of certain
keystone structures is sufficient to provide a high number of ant
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Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
FIGURE 2 | Effects of Log (Area) (A) and Connectivity (B) on Total number of occurrences estimated by the complete model (including the interaction; see Table 5).
Shaded regions represent 95% confidence intervals. The model was a GLM with Poisson distribution and log link function.
species (Hevia et al., 2013; e.g., grassland corridors in agricultural
landscapes: Azcárate et al., 2013). In other environments, such
as forest remains in urban matrices (Melliger et al., 2018) or
grassland remnants in forest matrices (Dauber et al., 2006), the
capacity of ants to maintain relatively rich communities in small
fragments has also been highlighted. Ants, therefore, would show
less vulnerability to fragmentation than other groups.
This apparently surprising ability to resist in small, relatively
isolated patches may be related to some ant features. Possibly,
the eusocial organization provides mechanisms to manage
the greater environmental stochasticity of small patches. For
example, most ant colonies are long-lived, and can maintain
food reserves that buffer fluctuations in resource availability
(Davidson, 1977), as well as regulate the size of their colonies, or
modify the task allocation of workers depending on conditions
(Gordon and Mehdiabadi, 1999). Ants are in general central-
place foragers (Harkness and Maroudas, 1985), so their resource-
catching strategy is based on the use of a particular territory, not
necessarily a large one (Azcárate and Peco, 2003). It is likely,
therefore, that the patch size threshold below which negative
population effects appear will be lower for ants than for other
animal groups such as bees, whose foraging depends on spatial
scales that can be very large (Greenleaf et al., 2007). Moreover,
most ant species produce winged sexuals, which gives them
a dispersal ability that may be enough to minimize isolation
between fragments (Dauber et al., 2006). Dispersal ability is often
complemented by other effective strategies to deal with habitat
patchiness, like poligyny or multiple queening (e.g., Plagiolepis,
Proformica, Solenopsis, Tapinoma) or flexible modalities of
nuptial flights (Bourke and Heinze, 1994). In addition, foraging
strategies may be completed in the apparently hostile matrix
(especially in the case of granivores) were weeds and crops can
provide some valuable resources (Baraibar et al., 2009, 2017).
Matrices in fragmented landscapes are not necessarily hostile
for all species, and in fact, management of the matrix has
often been pointed out as a key factor for the conservation of
the biodiversity within fragments (e.g., Armbrecht et al., 2001).
Lastly, resistance to habitat loss can be higher for generalist
species (Armbrecht et al., 2001; Marvier et al., 2004), and we
should not discard that the regional species pool is impoverished
since long ago, so that today it is dominated by generalist ants,
more tolerant to fragmentation. We think, however, that this
explanation is unlikely, given the elevated species densities found
in this study (higher than those observed in nearby locations,
Flores et al., 2018), and the fact that some of the species recorded
in our samplings are not so common in the Iberian context
(Temnothorax formosus, T. universitatis, and Goniomma blanci).
Particularly noteworthy is the unexpected negative effect
exerted by fragment area and connectivity on the total number
of occurrences per plot, a result that was also observed when
repeating the analysis with the total number of workers per
plot (Supplementary Table 1). Interestingly, this increase in
abundance for small and poorly connected patches was not
linked to any change in species composition, suggesting a general
effect favoring ants as a group. One possible explanation could
be that these remnants have a higher edge effect, which may
have more positive than negative consequences on ants (e.g.,
possibility of foraging in matrix habitats, higher productivity of
edges especially in agroecosystems where farmlands are fertilized,
etc.; Brascheler and Baur, 2003; Fahrig, 2017; Fahrig et al., 2019).
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Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
FIGURE 3 | Species richness estimated from rarefaction curves for each remnant. Species richness (horizontal lines) was estimated as the expected number of
species for 44 occurrences (vertical lines), which was the minimum number of occurrences observed in any of the 20 plots (F17). Occurrence-based rarefaction
curves (Gotelli and Colwell, 2011) were estimated by averaging 500 resamples with replacement. Bars are ±95%confidence intervals.
However, this explanation should be taken with caution since
plots were located at the same distance from the field edge (10 m)
regardless of fragment size. Other plausible explanation would
be related to the existence of refuge effects of small and isolated
patches against predators or parasites (Fahrig, 2017). In any
case, the observed effect would be limited to the range of sizes
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Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
TABLE 6 | Parameter estimates and P-values for the GLM (Gaussian distribution,
identity link function) built for species richness rarefied to 44 occurrences as a
function of Log (Area), Connectivity and the interaction Log (Area) x Connectivity.
Estimate Std. error t-value Pr(>|t|)
Intercept 16.404 3.261 5.030 0.0001
Log (Area) −1.196 0.782 −1.528 0.1460
Connectivity −10.589 6.848 −1.546 0.1416
Log (Area) ×Connectivity 2.390 1.722 1.388 0.1842
N=20 fragments.
TABLE 7 | Parameter estimates and P-values for the Permanova analysis built to
test the effect of area and connectivity on ant species community composition in
20 habitat remnants.
Df F. Model R2Pr(>F)
Area 1 0.79 0.04 0.553
Connectivity 1 2.01 0.09 0.129
Area ×Connectivity 1 2.51 0.12 0.063
Residuals 16 0.75
We used frequency data to estimate species composition and calculated dissimilarity
matrices based on Bray Curtis distance.
analyzed here, and we cannot exclude a hypothetical unimodal
response if smaller and more isolated remnants were included in
the study. Finally, it is also interesting that the increase in ant
abundance in small and isolated patches did not translate into an
increase in species density, suggesting that the rise in abundance
is experienced mostly by common species.
In the context of a generalized loss of habitat and
fragmentation exacerbation, protection of small and isolated
habitat remnants has traditionally been rarely considered in
favor of large patches (SLOSS debate), even though the same
amount of habitat spread over many small patches can result in
more biodiversity than when it is grouped into one or very few
large areas (Fahrig, 2017; Fahrig et al., 2019). Recently, however,
several studies have highlighted the important role of small
reserves and habitats in harboring biodiversity (Wintle et al.,
2019; Volenec and Dobson, 2020), which advises revisiting the
SLOSS debate. While there is no doubt that larger patches host
higher numbers of species in absolute terms, this effect is at least
in part a consequence of random sampling effects and species—
area relationships, and does not necessarily imply negative effects
attributable to fragmentation per se (Fahrig, 2017; Fahrig et al.,
2019), nor a higher conservation value for the area gathered
in large patches. There are also other mechanisms by which
habitat distribution in many smaller patches can provide positive
effects for biodiversity conservation and ecosystem services.
These mechanisms include the provision of spatial subsidies
by the matrix (Ewers and Didham, 2006), positive edge effects
(Ewers and Didham, 2006; Fletcher et al., 2018), landscape or
habitat complementarity (Fahrig, 2017) or a general increase
of beta diversity in heterogeneous landscapes (Andrén, 1994),
which ultimately means that fragmented landscapes often have
more diversity than non-fragmented ones when considered as
a whole. A higher fragmentation (therefore, a higher habitat
patchiness) can also produce positive effects at the population
level, by spreading the risk of extinction over a large number
of sites, thus reducing the risk of simultaneous extinction
of all local populations, stabilizing predator—prey or host—
parasitoid systems, or by reducing intraspecific and interspecific
competition (Fahrig, 2017). In addition, the existence of a high
number of small fragments in the territory can contribute to
overall landscape connectivity by acting as key “stepping stones”
for species able to move long distances (Herrera et al., 2017).
Finally, it should be noted that the small, poorly connected
patches of the dry agroecosystem studied here were not only
effective in maintaining ant communities, but also showed
high diversity values when compared to geographically close
areas. Thus, for example, Flores et al. (2018) analyzed ant
communities in grassland ecosystems of central Spain along a
wide environmental gradient using the same protocol as the
one followed here, and found that species density ranged from
2 to 15 species per plot, well below the 9–17 species per plot
found in our area. The cited study estimated the expected species
density as a function of altitude, obtaining values from 7 to
8 species per plot for the altitude corresponding to our study
area (700 m), also below the mean value of 11.8 found here.
Therefore, as with other groups (Escudero et al., 2015), the
high conservation value of gypsum ecosystems must also be
highlighted in the case of ants. Furthermore, this group plays
a central role in the functioning of terrestrial ecosystems, and
participates in the delivery of a number of ecological services
(Del Toro et al., 2012), so maintaining an appropriate mix of
species rich fragments interspersed with croplands is of particular
interest. In Mediterranean agroecosystems, the service of weed
control stands out, and is mainly performed by ants of the
genus Messor (Baraibar et al., 2009), which are very present
in the communities studied here. Other important services of
direct interest to agriculture are pest control, soil movement,
decomposition and nutrient cycling (Del Toro et al., 2012, 2015).
In summary, this study corroborates that the ecological
phenomena related to landscape configuration are complex and
can give rise to very different outcomes depending on the group
and the type of environment. It is therefore necessary to pay
attention to the particularities of each case, and to be aware of
the possible importance for biodiversity conservation of small
ecological fragments or structures, even though at first sight they
may not seem important, and especially when we are talking
about organisms that are valuable for the provision of ecosystem
services. This seems to be the case of ants in Mediterranean
gypsum agroecosystems, where we have observed that small and
weakly connected patches do not present losses of diversity with
respect to larger patches and higher connectivity. As a take home
message, we also want to recommend the maintenance of the
smallest remnants of natural habitats because they can shelter
important levels of biodiversity completing the list of services of
any agroecosystem.
DATA AVAILABILITY STATEMENT
The original contributions generated for the study are included
in the article/Supplementary Material, further inquiries can be
directed to the corresponding author/s.
Frontiers in Ecology and Evolution | www.frontiersin.org 8March 2021 | Volume 9 | Article 619215

Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
AUTHOR CONTRIBUTIONS
FA, AE, and AS planned and designed the research. AA-M,
FA, and AS conducted field work. AA-M and FA conducted
laboratory work. FA and AS conducted data statistical analyses.
FA wrote the manuscript with extensive input from the rest of
the authors. All authors contributed to the article and approved
the submitted version.
FUNDING
This study was supported by the Rey Juan Carlos University
(GYPVIFUN URJC, M2148), by the European Social Fund
managed by the Regional Government of Madrid (Remedinal
TE-CM: S2018/EMT-4338), and by the Horizon 2020 initiative
from the European Comission (Gypworld H2020-MSCA-RISE-
2007-777803).
ACKNOWLEDGMENTS
We thank Roberto López Rubio, Rufino Alameda Sánchez, and
Eugenio López Martínez for their help in fieldwork.
SUPPLEMENTARY MATERIAL
The Supplementary Material for this article can be found
online at: https://www.frontiersin.org/articles/10.3389/fevo.
2021.619215/full#supplementary-material
REFERENCES
Altieri, M. A. (1999). The ecological role of biodiversity in agroecosystems. Agric.
Ecosyst. Environ. 74, 19–31. doi: 10.1016/S0167-8809(99)00028-6
Andrén, H. (1994). Effects of habitat fragmentation on birds and mammals in
landscapes with different proportions of suitable habitat: a review. Oikos 71,
355–366. doi: 10.2307/3545823
Armbrecht, I., Tischer, I., and Chacón, P. (2001). Nested subsets and partition
patterns in ant assemblages (Hymenoptera,Formicidae) of Colombian dry
forest fragments. Pan-Pac. Entomol. 77, 196–209.
Assis, I. A., Dos Santos, F. N., Ramos, K. E., Barrios-Rojas, J. D., Majer,
J. D., and Vilela, E. F. (2018). Agricultural matrices affect ground ant
assemblage composition inside forest fragments. PLoS ONE 13:e0197697.
doi: 10.1371/journal.pone.0197697
Azcárate, F. M., and Peco, B. (2003). Spatial patterns of seed predation by harvester
ants (Messor Forel) in Mediterranean grassland and scrubland. Insect. Soc. 50,
120–126 doi: 10.1007/s00040-003-0635-y
Azcárate, F. M., and Peco, B. (2012). Abandonment of grazing in a Mediterranean
grassland area: consequences for ant assemblages. Insect Conserv. Divers. 5,
279–288. doi: 10.1111/j.1752-4598.2011.00165.x
Azcárate, F. M., Seoane, J., Castro, S., and Peco, B. (2013). Drove roads: keystone
structures that promote ant diversity in Mediterranean forest landscapes. Acta
Oecol. 49, 107–115. doi: 10.1016/j.actao.2013.03.011
Baraibar, B., Canadell, C., Torra, J., Royo-Esnal, A., and Recasens, J. (2017). Weed
seed fate during summer fallow: the importance of seed predation and seed
burial. Weed Sci. 65, 515–524. doi: 10.1614/WS-D-16-00031.1
Baraibar, B., Westerman, P. R., Carrión, E., and Recasens, J. (2009). Effects of tillage
and irrigation in cereal fields on weed seed removal by seed predators. J. Appl.
Ecol. 46, 380–387. doi: 10.1111/j.1365-2664.2009.01614.x
Barto´
n, K. (2009). MuMIn: multi-model inference. R Package, Version 0.12.12.
Bourke, A. F. G., and Heinze, J. (1994). The ecology of communal breeding:
the case of multiple-queen Leptothoracine ants. Philos. Trans. Biol. Sci. 345,
359–372 doi: 10.1098/rstb.1994.0115
Brascheler, B., and Baur, B. (2003). Effects of experimental small-scale grassland
fragmentation on spatial distribution, density, and persistence of ant nests. Ecol.
Entomol. 28, 651–658 doi: 10.1111/j.1365-2311.2003.00549.x
Colwell, R. K. (2013). EstimateS: statistical estimation of species richness and
shared species from samples. Version 9.1.0.
Comas, C., Royo-Esnal, A., Recasens, J., and Torra, J. (2016). Analysing spatial
correlation of weeds and harvester ants in cereal fields using point processes.
Arthropod Plant Interact. 10, 197–205. doi: 10.1007/s11829-016-9425-0
Concostrina-Zubiri, L., Martínez, I., and Escudero, A. (2018). Lichen-
biocrust diversity in a fragmented dryland: fine scale factors are better
predictors than landscape structure. Sci. Total Environ. 628–629, 882–892.
doi: 10.1016/j.scitotenv.2018.02.090
Dauber, J., Bengtsson, J. A. N., and Lenoir, L. (2006). Evaluating effects of habitat
loss and land-use continuity on ant species richness in seminatural grassland
remnants. Conserv. Biol. 20, 1150–1160. doi: 10.1111/j.1523-1739.2006.00373.x
Davidson, D. W. (1977). Ecology and community organization in desert seed-
eating ants. Ecology 58, 725–737. doi: 10.2307/1936209
Del Toro, I., Ribbons, R. R., and Ellison, A. M. (2015). Ant-mediated ecosystem
functions on a warmer planet: effects on soil movement, decomposition and
nutrient cycling. J. Anim. Ecol. 84, 1233–1241. doi: 10.1111/1365-2656.12367
Del Toro, I., Ribbons, R. R., and Pelini, S. L. (2012). The little things that run the
world revisited: a review of ant-mediated ecosystem services and disservices
(Hymenoptera: Formicidae). Myrmecol. News 17, 133–146.
Emmerson, M., Morales, M. B., Oñate, J. J., Batary, P., Berendse, F., Liira, J.,
et al. (2016). How agricultural intensification affects biodiversity and ecosystem
services. Adv. Ecol. Res. 55, 43–97. doi: 10.1016/bs.aecr.2016.08.005
Escavy, J. I., Herrero, M. J., and Arribas, M. E. (2012). Gypsum resources
of Spain: temporal and spatial distribution. Ore geol. Rev. 49, 72–84.
doi: 10.1016/j.oregeorev.2012.09.001
Escudero, A., Palacio, S., Maestre, F. T., and Luzuriaga, A. L. (2015). Plant
life on gypsum: a review of its multiple facets. Biol. Rev. 90, 1–18.
doi: 10.1111/brv.12092
Ewers, R. M., and Didham, R. K. (2006). Confounding factors in the detection
of species responses to habitat fragmentation. Biol. Rev. 81, 117–142.
doi: 10.1017/S1464793105006949
Fahrig, L. (2003). Effects of habitat fragmentation on biodiversity. Annu. Rev. Ecol.
Evol. Syst. 34:487–515. doi: 10.1146/annurev.ecolsys.34.011802.132419
Fahrig, L. (2017). Ecological responses to habitat fragmentation per se.Annu. Rev.
Ecol. Evol. Syst. 48, 1–23. doi: 10.1146/annurev-ecolsys-110316-022612
Fahrig, L., Arroyo-Rodríguez, V., Bennett, J. R., Boucher-Lalonde, V., Cazetta, E.,
Currie, D. J., et al. (2019). Is habitat fragmentation bad for biodiversity? Biol.
Conserv. 230, 179–186. doi: 10.1016/j.biocon.2018.12.026
Fletcher, R. J., Didham, R. K., Banks-Leite, C., Barlow, J., Ewers, R. M., Rosindell,
J., et al. (2018). Is habitat fragmentation good for biodiversity? Biol. Conserv.
226, 9–15. doi: 10.1016/j.biocon.2018.07.022
Flores, O., Seoane, J., Hevia, V., and Azcárate, F. M. (2018). Spatial patterns
of species richness and nestedness in ant assemblages along an elevational
gradient in a Mediterranean mountain range. PLoS ONE 13:e0204787.
doi: 10.1371/journal.pone.0204787
Folgarait, P. (1998). Ant biodiversity and its relationship to ecosystem functioning:
a review. Biodivers. Conserv. 7, 1221–1244. doi: 10.1023/A:1008891901953
Golden, D. M., and Crist, T. O. (2000). Experimental effects of habitat
fragmentation on rove beetles and ants: patch area or edge? OIKOS 90, 525–538.
doi: 10.1034/j.1600-0706.2000.900311.x
Gordon, D. M., and Mehdiabadi, N. J. (1999). Encounter rate and task allocation
in harvester ants. Behav. Ecol. Sociobiol. 45, 370–377. doi: 10.1007/s0026500
50573
Gotelli, N. J., and Colwell, R. K. (2001). Quantifying biodiversity: procedures and
pitfalls in the measurement and comparison of species richness. Ecol. Lett. 4,
379–391. doi: 10.1046/j.1461-0248.2001.00230.x
Gotelli, N. J., and Colwell, R. K. (2011). “Estimating species richness,” in Biological
Diversity: Frontiers in Measurement and Assessment, eds A. E. Magurran and B.
J. McGill (Oxford: Oxford University Press), 39–54.
Frontiers in Ecology and Evolution | www.frontiersin.org 9March 2021 | Volume 9 | Article 619215

Azcárate et al. Ants in Gypsum Mediterranean Agroecosystem
Greenleaf, S. S., Williams, N. M., Winfree, R., and Kremen, C. (2007). Bee
foraging ranges and their relationship to body size. Oecologia 153, 589–596.
doi: 10.1007/s00442-007-0752-9
Haddad, N. M., Brudvig, L. A., Clobert, J., Davies, K. F., Gonzalez, A., Holt, R. D.,
et al. (2015). Habitat fragmentation and its lasting impact on Earth’s ecosystems.
Appl. Ecol. 1, 1–9. doi: 10.1126/sciadv.1500052
Harkness, R. D., and Maroudas, N. G. (1985). Central place foraging by an ant
(Cataglyphis bicolor Fab.): a model of searching. Anim. Behav. 33, 916–928.
doi: 10.1016/S0003-3472(85)80026-9
Herrera, L. P., Sabatino, M. C., Jaimes, F. R., and Saura, S. (2017). Landscape
connectivity and the role of small habitat patches as stepping stones: an
assessment of the grassland biome in South America. Biodivers. Conserv. 26,
3465–3479. doi: 10.1007/s10531-017-1416-7
Herrero, J., and Porta, J. (2000). The terminology and the concepts of
gypsum-rich soils. Geoderma 96, 47–61. doi: 10.1016/S0016-7061(00)
00003-3
Hevia, V., Azcárate, F. M., Oteros-Rozas, E., and González, J. A. (2013).
Exploring the role of transhumance drove roads on the conservation of ant
diversity in Mediterranean agroecosystems. Biodivers. Conserv. 22, 2567–2581.
doi: 10.1007/s10531-013-0539-8
Kennedy, C. M., Lonsdorf, E., Neel, M. C., Williams, N. M., Ricketts, T. H.,
Winfree, R., et al. (2013). A global quantitative synthesis of local and landscape
effects on wild bee pollinators in agroecosystems. Ecol. Lett. 16, 584–599.
doi: 10.1111/ele.12082
Levins, R. (1969). Some demographic and genetic consequences of environmental
heterogeneity for biological control. Bulletin of the Entomol. Soc. Am. 15,
237–240. doi: 10.1093/besa/15.3.237
Longino, J. T., Coddington, J., and Colwell, R. K. (2002). The ant fauna of a tropical
rain forest: estimating species richness three different ways. Ecology 83, 689–702
doi: 10.1890/0012-9658(2002)083[0689:TAFOAT]2.0.CO;2
Luzuriaga, A. L., Sánchez, A. M., López-Angulo, J., and Escudero, A.
(2018). Habitat fragmentation determines diversity of annual plant
communities at landscape and fine spatial scales. Basic Appl. Ecol. 29,
12–19. doi: 10.1016/j.baae.2018.03.008
Marvier, M.,., Kareiva, P., and Neubert, M. G. (2004). Habitat
destruction, fragmentation, and disturbance promote invasion
by habitat generalists in a multispecies metapopulation.
Risk Anal. 24, 869–878. doi: 10.1111/j.0272-4332.2004.
00485.x
Matesanz, S., Escudero, A., and Valladares, F. (2009). Impact of three
global change drivers on a Mediterranean shrub. Ecology 90, 2609–2621.
doi: 10.1890/08-1558.1
Matesanz, S., Gómez-Fernández, A., Alcocer, I., and Escudero, A. (2015).
Fragment size does not matter when you are well connected: effects of
fragmentation on fitness of coexisting gypsophiles. Plant Biol. 17, 1047–1056.
doi: 10.1111/plb.12329
Melliger, R. L., Braschler, B., Rusterholz, H.-P., and Baur, B. (2018).
Diverse effects of degree of urbanisation and forest size on species
richness and functional diversity of plants, and ground surface-active
ants and spiders. PLoS ONE 13:e0199245. doi: 10.1371/journal.pone.
0199245
Moncalvillo, B., Matesanz, S., Escudero, A., and Sánchez, A. M. (2021). Habitat
fragmentation and population features differently affect fruit predation,
fecundity and offspring performance in a non-specialist gypsum plant. Plant
Biol. 23, 184–192. doi: 10.1111/plb.13183
Oksanen, J., Blanchet, F. G., Kindt, R., Legendre, P., Minchin, P. R., O’Hara, R.,
et al. (2015). Vegan: community ecology package. R Package Version 2.3-2.
Available onlibe at: http://CRAN.R- project.org/package=vegan
QGIS.org (2018). QGIS Geographic Information System. Open Source Geospatial
Foundation Project. Available onlibe at: http://qgis.org
Rabasa, S., Gutiérrez, D., Escudero, A., and Spence, J. (2005). Egg laying by a
butterfly on a fragmented host-plant: a multi-level approach. Ecography 28,
629–639. doi: 10.1111/j.2005.0906-7590.04229.x
Rabasa, S. G., Gutiérrez, D., and Escudero, A. (2009). Temporal variation in the
effects of habitat fragmentation on reproduction of the Mediterranean shrub
Colutea hispanica.Plant Ecol. 200, 241–254. doi: 10.1007/s11258-008-9448-4
Santamaría, S., Sánchez, A. M., López-Angulo, J., Ornosa, C., and Escudero, A.
(2018). Landscape effects on pollination networks in Mediterranean gypsum
islands. Plant Biol. 20, 184–194. doi: 10.1111/plb.12602
Santos, L. A. O., Bischoff, A., and Fernandes, O. A. (2018). The effect of forest
fragments on abundance, diversity and species composition of predatory ants
in sugarcane fields. Basic Appl. Ecol. 33, 58–65. doi: 10.1016/j.baae.2018.08.009
Silvestre, M., Aguilar, A., Seoane, J., and Azcárate, F. M. (2019). Responses of
seed size, ant worker size, and seed removal rate to elevation in Mediterranean
grasslands. Oecologia 189, 781–793. doi: 10.1007/s00442-019-04356-6
Tremlová, K., and Münzbergová, Z. (2007). Importance of species traits
for species distribution in fragmented landscape. Ecology 88, 965–977.
doi: 10.1890/06-0924
Tscharntke, T., Sekercioglu, C. H., Dietsch, T. V., Sodhi,N. S., Hoehn, P.,
and Tylianakis, J. M. (2008). Landscape constraints on functional diversity
of birds and insects in tropical agroecosystems. Ecology 89, 944–951.
doi: 10.1890/07-0455.1
Villard, M. A., and Metzger, J. P. (2014). Beyond the fragmentation debate: a
conceptual model to predict when habitat configuration really matters. J. Appl.
Ecol. 51, 309–318. doi: 10.1111/1365-2664.12190
Volenec, Z. M., and Dobson, A. P. (2020). Conservation value of small reserves.
Conserv. Biol. 34, 66–79. doi: 10.1111/cobi.13308
Whittaker, R. J., Willis, K. J., and Field, R. (2001). Scale and species richness:
towards a general, hierarchical theory of species diversity. J. Biogeogr. 28,
453–470. doi: 10.1046/j.1365-2699.2001.00563.x
Wintle, B. A., Kujala, H., Whitehead, A., Cameron, A., Veloz, S., Kukkala,
A., et al. (2019). Global synthesis of conservation studies reveals the
importance of small habitat patches for biodiversity. PNAS 116, 909–914.
doi: 10.1073/pnas.1813051115
Conflict of Interest: The authors declare that the research was conducted in the
absence of any commercial or financial relationships that could be construed as a
potential conflict of interest.
Copyright © 2021 Azcárate, Alameda-Martín, Escudero and Sánchez. This is an
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