Urbanisation and habitat fragmentation favour thermophilic and monogynous ant species

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Skip to main content Home About Submit ALERTS / RSS Search for this keyword Advanced Search New Results Urbanisation and habitat fragmentation favour thermophilic and monogynous ant species View ORCID Profile Basile Finand , View ORCID Profile Thibaud Monnin , Céline Bocquet , Angélique Bultelle , View ORCID Profile Pierre Fédérici , Léa Darmedru , Solène Gouffault , Joséphine Ledamoisel , View ORCID Profile Nicolas Loeuille doi: https://doi.org/10.1101/2025.02.25.640022 Basile Finand 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France 2 Ecosystems and Environment Research Program, Faculty of Biological and Environmental Sciences, University of Helsinki , Helsinki, Finland Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Basile Finand For correspondence: finand.basile{at}gmail.com Thibaud Monnin 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Thibaud Monnin Céline Bocquet 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site Angélique Bultelle 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site Pierre Fédérici 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Pierre Fédérici Léa Darmedru 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site Solène Gouffault 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site Joséphine Ledamoisel 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site Nicolas Loeuille 1 Sorbonne Université, Université Paris Cité, Université Paris Est Créteil, CNRS, INRAE, IRD, Institut d’Ecologie et des Sciences de l’Environnement de Paris (UMR7618) , 75005 Paris, France Find this author on Google Scholar Find this author on PubMed Search for this author on this site ORCID record for Nicolas Loeuille Abstract Full Text Info/History Metrics Supplementary material Data/Code Preview PDF Abstract Environmental changes such as urbanisation and habitat fragmentation profoundly impact ecological communities by altering habitats, resources, and microclimate. Ants, with diverse life histories and strong ecological effects, are ideal models to study these pressures. We investigated the response of ant communities, including taxonomic and functional diversity, to urbanisation and habitat fragmentation in the Paris region, comparing 24 urban parks vs 25 rural forests outside the city. We found a clear difference in species composition between urban and rural environments, with a higher prevalence of monogynous and thermophilic species in the city. Forest communities are homogeneous across the three fragmentation levels we studied, while park communities differ noticeably depending on park size, with larger parks harbouring more species. Our findings suggest that urbanisation selects specific ant traits and favours more thermophilic species, thereby increasing the mean thermal preference of urban communities. These selective effects influence which species can colonise and survive in different patches, shaping metacommunity structure and potentially affecting the resilience of ant communities under climate change. Introduction Habitat fragmentation stands as a primary driver of biodiversity decline, exerting manifold impacts on ecosystems worldwide ( Alberti, 2015 ; Aronson et al., 2014 ; Diamond & Martin, 2021 ; Haddad et al., 2015 ; Johnson & Munshi-South, 2017 ; Piano et al., 2020 ). This phenomenon is characterised by a reduction in total habitat size, with an increase in patch number, but smaller and more isolated ( Fahrig, 2003 ). Habitat fragmentation impedes dispersal between populations, leading to direct impacts on individual and gene flows, resulting in population declines and reductions in genetic variability ( Hastings, 1983 ; Travis et al., 2013 ; Young et al., 1996 ). These effects reverberate through ecological communities, altering their composition and dynamics. Habitat fragmentation gives rise to metacommunities, i.e. interconnected communities linked through dispersal. Metacommunity dynamics can happen in various ways depending on species and environmental characteristics ( Leibold et al., 2004 ). In scenarios where variations in environmental conditions among patches do not play a key role, biodiversity hinges exclusively on the competitive abilities and colonisation capacities of species, a dynamic termed patch dynamic. Colonising species can exploit and establish populations in patches where competitors cannot persist ( Tilman, 1994 ). Consequently, habitat fragmentation, through habitat destruction and isolation, is expected to favour colonising species while disadvantaging more competitive species, thereby reducing total diversity ( Finand, Monnin, et al., 2024 ; Tilman et al., 1994 ). Conversely, in environments where variability among patches drives ecological dynamics, two dispersal-dependent mechanisms emerge. In scenarios with restricted dispersal, species diversity is governed by species sorting, where species whose ecological niches best match local conditions competitively exclude others within each patch. Conversely, when dispersal between patches is extensive, mass effects come into play ( Mouquet & Loreau, 2003 ), allowing less-adapted species to persist in small numbers, being continually resupplied from surrounding patches, akin to source-sink dynamics. Although species sorting typically yields lower species richness per patch compared to mass effect, when dispersal is exceedingly high and mass effects dominate, all species compete for average metacommunity conditions, leading to competitive exclusion at this scale. Diversity is then low, at the local and metacommunity scale ( Mouquet & Loreau, 2003 ). The heightened fragmentation of habitat, by augmenting patch isolation, is expected to foster species sorting, consequently diminishing local species richness and increasing species turnover among patches. Finally, the neutral theory, akin to island biogeography ( MacArthur & Wilson, 1967 ), posits that diversity is solely dependent on species loss through emigration and extinction balanced by gains through immigration and speciation, with no differentiation in species abilities ( Hubbell, 2001 ). Importantly, these four paradigms can be disentangled based on variations in species composition and patterns of local (alpha), regional (gamma) diversity and variations in species composition among patches (beta diversity). For instance, if species sorting dominates, local composition should differ in contrasted environments so that beta diversity is mostly associated with a “turnover” component (sensu ( Baselga, 2010 )). Conversely, local dispersal in mass effects should lead to a spatial autocorrelation in species composition so that communities in harsher environments or smaller patches should be a subset of communities developing in better or larger patches. This enhances the “nestedness” component of beta diversity (sensu ( Baselga, 2010 )). Urbanisation is another major factor affecting biodiversity. This transformative process reshapes environments in many ways, fostering novel ecological dynamics and posing significant challenges to native flora and fauna. First, urbanisation is most often associated with high habitat fragmentation ( Haddad et al., 2015 ). Second, it also generates varied types of pollution, including chemicals, noise disturbance, and altered light regimes, often favouring species with specific adaptations or tolerances ( Hölker et al., 2010 ; Newport et al., 2014 ). These pollutants permeate air, water, and soil, exerting deleterious effects on biotic components of ecosystems. Third, changes in resource availability represent another hallmark of urbanisation, including variations in quantity, quality, and seasonality ( Hantak et al., 2021 ; Wilson & Jamieson, 2019 ). Finally, the creation of distinct microclimates within cities, characterised by elevated temperatures (urban heat island effect), exacerbates thermal stress on organisms and modifies ecosystem dynamics ( Kaiser et al., 2016 ; McGlynn et al., 2019 ). This thermal gradient can influence species distributions, phenology, and physiological processes, driving community composition and structure shifts. Beyond diversity indices, the specific environmental factors and the low connectivity encountered in cities should select specific traits. Habitat fragmentation may favour low dispersal ( Cheptou et al., 2008 ; Finand et al., 2023 ), while higher temperatures promote species and phenotypes whose thermal preference matches these conditions ( Campbell-Staton et al., 2021 ; Diamond et al., 2018 ). For instance, longer-legged species are selected in hotter parks in Taichung City because the distance between the body and soil limits desiccation ( Liu et al., 2019 ). In general, urban and rural species may differ in many other traits ( Buchholz & Egerer, 2020 ; Croci et al., 2008 ) reflecting the multidimensional changes in such environments. Insects and in particular ants have an important impact on ecosystems by dispersing seeds, pollinating, participating in nutrient cycles, structuring the soil, and participating in food chains ( Hölldobler & Wilson, 1990 ). They are a very good model group to study the effects of urbanisation and habitat fragmentation on communities. First, their life history traits are variable between species. Some of these traits are strongly related to their dispersal abilities (size of individuals, presence or absence of wings in queens) and competitiveness (size of colonies, diet), and therefore play a key role as functional traits in metacommunity dynamics. Second, ants are abundantly present and diverse in and outside cities, which allows direct comparative studies. Most ant species depend on green spaces, which makes the city a very fragmented habitat compared to the countryside, even if some species can live in built areas. Last, they are easy to sample. Several empirical studies have shown that urbanisation and habitat fragmentation affect the species richness of ants, with a decrease in species richness as urbanisation increases and a change in community composition ( Carpintero & Reyes-López, 2014 ; Ješovnik & Bujan, 2021 ; Leal et al., 2012 ; Pacheco & Vasconcelos, 2007 ; Theunis et al., 2005 ). For instance, climate and resource specialist ants are more sensitive to habitat fragmentation ( Leal et al., 2012 ). Moreover, this change in community composition due to habitat fragmentation impacts ecosystem services provided by ants influencing decomposition and soil structure ( Sanford et al., 2009 ). Paris and its region offer a good opportunity to study urbanisation and habitat fragmentation, with a marked gradient from the very densely populated city centre to low-density areas thirty kilometres away. The city has numerous urban parks of various sizes, surrounded by a matrix of streets and buildings unsuitable for most ant species. This makes Parisian parks a perfect setup to understand the impact of fragmentation on ant communities. Moreover, the urban heat island effect within Paris is well-documented ( Lemonsu et al., 2015 ). The surrounding region is more rural with many natural forests of various sizes allowing an interesting comparison with the city. We here investigate the impact of urbanisation and habitat fragmentation on ant communities at different scales by comparing forest (low urbanisation and low fragmentation) and urban communities (high urbanisation and high fragmentation). We measured community differences between the two habitats, and the effects of the two main components of fragmentation, isolation, and patch size, within each habitat. We hypothesise that fragmentation decreases species richness and changes community structure, with more turnover than nestedness due to higher species sorting and lower mass effects. Furthermore, we studied variations in traits linked to thermal preference, colony social structure (queen number) and dispersal abilities. We hypothesise that the city harbours species that are more thermal-resistant due to the urban heat island effect, and that have short dispersal distances because of the high habitat fragmentation. Material and methods Patch selection To study the impact of urbanisation and particularly of habitat fragmentation on litter ants, we sampled two different habitats: 25 rural forests in the Paris region and 24 wooded parts of urban parks within the city ( Figure 1 ). Forests are considered minimally fragmented environments, as they offer a favourable habitat for litter ants and are surrounded by a matrix of habitats that, while less favourable, can still support ant species, e.g. hedges around agricultural fields or private gardens. On the contrary, parks are islands of favourable habitats (woods and shrubs) surrounded by a mostly hostile matrix (streets and buildings), though local refugia may exist. To compare the same litter ant communities in forests and parks, we targeted park areas with natural litter for sampling, i.e., areas under trees or shrubs. This allowed us to focus on the specific community of ants living in the litter and to make habitats as comparable as possible. Evidence that we sampled the forest litter ant community in parks is that the two species most abundant in our park samples are Temnothorax nylanderi and Myrmecina graminicola , which are particularly abundant in forests ( Blatrix et al., 2013 ; Seifert, 2018 ) and also the most abundant ants in our forest samples (see results). Forests adjacent to the city are used here as reference communities to assess the effect of urbanisation, which intrinsically involves habitat change, as it is the closest environment to litter. Differences between park and forest habitats reflect spatial variations in niche conditions, which is precisely what this study aims to examine. We study the effect of urbanisation at the regional scale (city parks vs forests) and the effect of habitat fragmentation at the local scale (parks of various sizes and isolation, forests of various sizes). Download figure Open in new tab Figure 1: Map of the sampling sites in forests around Paris (Ile-de-France region) (a) and parks in Paris (b). The colour of the circles indicates the fragmentation level of forests or the size of parks. minimally fragmented forests (n = 10) have between 75 and 95% of forest cover within 1 Km, medium fragmented forests (n = 8) have between 30 and 70% of forest cover, and highly fragmented forests (n = 6) have less than 5% of forest cover in their surroundings. Parks are divided based on surface: large (80 000 to 250 000 m 2 (n = 8)), medium (5 000 to 12 000 m 2 (n = 9)), and small (500-2 000 m 2 (n = 8)). Site information is detailed in Table S1. We selected forest patches (n = 24) of three fragmentation classes ( Figure 1a ). Minimally fragmented forests (n = 10) had between 75 and 95% of forest cover within 1 Km, medium fragmented forests (n = 8) had between 30 and 70% of forest cover, and highly fragmented forests (n = 6) had less than 5% of forest cover in their surroundings. We used the database BD Forêt® Version 2.0 from the “National Institute of Geographic and Forest Information” of France (IGN, https://inventaire-forestier.ign.fr/carto/afficherCarto/V2 , last visit: 19/07/2022) to assess forest cover. Forest classes are evenly distributed around Paris, and we chose well-separated locations that likely represent independent plots (mean distance = 52.1 +/- 23.7 Km, min = 6.4 Km, max = 122.1 Km). Parks fragmentation was measured using two metrics, size and isolation. We selected parks (n = 25) of three size classes ( Figure 1b ), namely large (n = 8, 80 000-250 000 m 2 ), medium (n = 9, 5 000-12 000 m 2 ), and small parks (n = 8, 500-2 000 m 2 ), and measured their isolation as the quantity of green area (other parks) within 1 Km around the focal park boundaries. As for forests, we selected parks that were evenly distributed within Paris and well separated, assumed to represent independent points (mean distance = 4.9 +/- 2.1 Km, min = 552 m, max = 9.5 Km). Sampling We sampled forests between May and July 2020, between 11 am and 4 pm, on rainless days. We distributed 12 quadrats of one square meter each, equally spaced every 9 m, along a transect of 100 m. We independently collected the litter and superficial soil on each of the 12 quadrats. We collected organisms using the Winkler extraction method ( Agosti et al., 2000 ). We detail the procedure below. We sampled parks between May and August 2021, between 11 am and 4 pm, on rainless days. To compare the same litter ant communities in forests and parks, we targeted park areas with natural litter for sampling, i.e. areas under trees or shrubs. It allows us to focus on the specific community of ants living in the litter and make the habitats as comparable as possible with forests. We purposefully placed the 12 quadrats into these areas and dispersed the quadrats within parks as much as possible (considering the configuration of the parks, transect was not possible). One park (OLB) was so small that only five quadrats could be placed and sampled, and owing to a disturbance only ten quadrats were sampled in another park (MAS). The sampling and extraction protocols were the same for parks and forests. For each quadrat, we sifted the litter and the first centimetres of soil using a 1 cm 2 sieve (Figure S1a). This removed leaves, twigs and stones and retained the soil and organisms within, including ants. We kept what passed the sieve in large plastic bags until we returned to the laboratory. There, we moved the litter to sieve bags that were hung in a Winkler extractor for 48 hours ( Agosti et al., 2000 ) (Figure S1b). As the litter gradually dried out, organisms moved away and fell into an alcohol-filled collecting vial. This method samples all the macrofauna living in the collected soil and litter, including ant foragers present in the litter and parts of the colonies that nest in the litter or superficially in the soil. We sampled litter ants in forests and parks in two separate years (2020 and 2021) due to logistical constraints and movement restrictions related to the COVID-19 pandemic. Note that ant queens and colonies live several years, and that the sampling was done during the same season. While fluctuations of abundances from one year to another are certainly possible, our analyses focus on species presence/absence in each quadrat rather than abundance. We further discuss the implications of our sampling scheme in the discussion section. Identification of ant species We identified ants at the species level using a binocular microscope and three identification keys ( Blatrix et al., 2013 ; Seifert, 2018 ); Galkowski & Lebas “Identification des Myrmica ”, unpublished). In addition, Temnothorax and Hypoponera species were identified by Xavier Espadaler, while Tetramorium species were identified by Bernard Kaufmann. Ant species traits We selected different species-level traits to test for the effect of urbanisation. Queen size, worker size and colony size (number of individuals) were obtained from the bibliography for all collected species ( Blatrix et al., 2013 ; Seifert, 2018 ). In addition, colony mass was estimated as worker size 3 X colony size. Because colony mass is directly linked to metabolic requirements ( Brown et al., 2004 ), we expect that it could vary between urban and forest sites, either due to changes in resources or temperature regimes. We also analysed the social structure of a colony, with two levels: monogyny (species with typically one queen per colony) and polygyny (species with typically several queens per colony). This trait is to some extent linked to dispersal as monogynous species often perform independent colony foundation (long dispersal distance strategy but low competitivity) whereas polygynous species do colony fission (short dispersal distance strategy but high competitivity) ( Cronin et al., 2013 ; Keller, 1991 ). Finally, for each species, we estimated thermophily based on a meridional index from Daufresne et al. (2004) ( Daufresne et al., 2004 ). Recent advances in ant thermal ecology have highlighted more precise traits such as critical thermal limits, activity temperatures, or behavioural tolerances ( Nascimento et al., 2022 ; Parr & Bishop, 2022 ; Roeder et al., 2021 ). They are well-suited to characterise precisely the thermal performances of a few species in a given environment. However, they require live workers in significant numbers (20 to 30 workers per colony, more for polymorphic species, and several colonies per species), making them incompatible with our sampling protocol with Winklers. The meridional index is better suited for analysing diverse assemblages across broad spatial scales, like in our study. In addition, it integrates factors other than thermal resistance per se that may relate to thermal preference, e.g. circadian rhythm of activity, and it is generalisable across taxa and compatible with studies on biogeographical shifts under climate change. It provides a useful, scalable metric for detecting broad patterns of community shifts along environmental gradients. Species were considered meridional if most of their distribution lies south of Paris, with meridionality being calculated by (Paris latitude – minimum observed latitude) / (maximum observed latitude – minimum observed latitude). The meridional index for a given species is, therefore, close to one when Paris is at the northern limit of its distribution, and close to zero when Paris is at the southern limit of its distribution. Minimum and maximum latitudes were obtained for each species distribution from the antmap.org website (last visit: 28/02/2024; ( Guénard et al., 2017 ; Janicki et al., 2016 )). Only the records considered to pertain to the native distribution were considered. Following previous works, we use this index as a proxy of the thermal preference of a species. A Parisian species at the northern limit of its distribution is more thermophilic than one at the southern limit. Paris is farther north than the northern range limit for two species. We have attributed 1, the highest value, to them. To understand which traits are linked to urbanisation, we assign a value of urbanisation preference for each species, calculated as the number of urban patches where the species has been found divided by the total number of patches (urban and forest patches) where it has been found. Strictly urban and forest species have urbanisation indices of one and zero, respectively. Analyses Statistical analyses were done with R v3.6.2 (R Core Team 2016). We checked that our sampling was sufficient to sample the litter ant biodiversity with rarefaction curves and chao analysis using the vegan package. We studied alpha diversity as a function of urbanisation at the regional scale (forests vs. parks) using ANOVAs. At the local scale (among parks and among forests), the observed species richness and the Shannon index were analysed as a function of park size and isolation, and forest fragmentation using ANOVAs. We checked that the data complied with application conditions using the DHARMa package ( Hartig, 2022 ), and log-transformed them when normality or homoscedasticity were not satisfied. We analysed beta diversity with Jaccard’s dissimilarity index because we here focus on species presence/absence rather than on abundance. Direct use of individual densities could be problematic, as our sampling procedure can recover isolated individuals as well as whole colonies for some Temnothorax species and Myrmecina graminicola . The second instance would inflate recorded abundances of superficial species compared to species nesting deeper in the soil for which we only collect foragers. We performed db-RDA (distance-based redundancy analysis), an RDA on dissimilarity distances, to assess if our environmental factors affect dissimilarity between our communities ( Legendre & Anderson, 1999 ). We used ANOVAs to test the significance using the vegan package. We used the package betapart to assess the percentage of turnover and nestedness composing the dissimilarities between our communities ( Baselga & Orme, 2012 ). To link the urbanisation preference to species traits, we first checked the correlation among traits using a Pearson test. The different size measurements were highly correlated (Figure S2). We kept the total mass of a colony, as it contains the size of a colony and the weight of individuals. In the end, variations in three traits were explored: the total mass of a colony, the social structure, and the meridional index. To assess if the urbanisation preference of species is correlated with each trait, we performed a linear model for the colony mass and the meridional index and a generalised linear model with a binomial link function for the social structure. We tested the application conditions (normality, homoscedasticity) using the DHARMa package ( Hartig, 2022 ). We log-transformed the total mass. Results We collected and identified 36 361 ants from 12 genera and 29 species. Abundances are similar in forests and parks (18 094 and 18 267 ants, respectively). Similarly, the number of species recovered per habitat type is comparable ( Figure 2 , 23 species in forests and 20 species in parks). 14 species occurred in both parks and forests, albeit in most cases with different relative abundances, while 9 species occurred in forests and 6 in parks only ( Figure 2 ). One invasive species only was recorded ( Lasius neglectus ). Removing it from the analyses does not qualitatively change the results. Download figure Open in new tab Figure 2: Variations in the presence and abundance of ant species depending on patches. Darker Levels show higher abundances, i.e., of the percentage of quadrats where the species was found. On the Y-axis, green patches correspond to forests and grey patches correspond to parks. Red, orange, and yellow show fragmentation levels for forests and patch sizes for parks. Species are ranked based on their urbanisation preference along the X-axis. Sampling effort The rarefaction curves and the chao analyses show that the sampling was sufficient to estimate litter ant diversity (Figure S3). Indeed, we sampled 95.19 +/- 25.41 % of the theoretical number of species in forests and 91.13 +/- 13.12 % in parks. Only two forests (CSM and GIF) and three parks (HEA, HEG and PDG) have a percentage of observed species representing less than 80% of the theoretical number of species. However, this theoretical number of species is higher in these patches (around or above 15) than in the other patches (usually around 10, seldom higher than 12) so it may be overestimated. Therefore, we use the observed number of species per patch in the analyses. Urbanisation effect at the regional scale (forests vs parks) Unexpectedly, we found more species locally with increased urbanisation (F value = 6.67, p = 0.013, R 2 = 10.6%, Figure 3a ). Indeed, parks support 8.96 +/- 1.74 species while forest sites have 7.58 +/- 2.30 species. In agreement with this, the Shannon index is higher in the city environment showing a higher diversity with less dominant species (T value = 2.808, p = 0.007, R 2 = 13%, Figure 3d ). Download figure Open in new tab Figure 3: Species richness (a-c) and Shannon index (d-f) between forests and parks (a, d), between parks of different sizes (b, e), and between forests of different fragmentation levels (c, f). Significant differences are observed in species richness and Shannon index between forests and parks (Species richness: F value = 6.67, p = 0.013, R 2 = 10.6%; Shannon index: T value = 2.808, p = 0.007, R 2 = 13%), and in species richness between parks of different sizes (F = 3.74, p = 0.039, R 2 = 0.19). Urbanisation also influences species composition (F value = 28.47, p = 0.001, R 2 = 37.7%, Figure 4a ). Communities are clearly different between parks and forests, with 89% of the beta diversity due to turnover (Jaccard index = 0.517, Figure 2 ), suggesting a dominant role of species sorting. Download figure Open in new tab Figure 4: Beta diversity (db-RDA analysis) between forests and parks (a), between parks (b), and between forests (c). Blue acronyms are the species. Capital acronyms are the sites. For (c), yellow represents high fragmentation, orange medium fragmentation, and red low fragmentation. Ant communities are markedly different between parks and forests, with communities from the two habitats being separated with no overlap (F value = 28.47, p = 0.001, R 2 = 37.7%). Communities also clearly differ between parks of different sizes, with a gradual change in composition from small to medium to large parks (F value = 2.14, p = 0.008, R 2 = 16.3%). No difference is observed between forests of different fragmentation level (F = 0.77, p = 0.69). Fragmentation effect at the local scale The fragmentation level at the local scale has no effect. In forests, it does not affect species richness (F value = 0.17, p = 0.84, Figure 3c ), Shannon index (F value = 0.488, p = 0.62, Figure 3f ) or the composition of ant communities (F value = 1.37, p = 0.128, Figure 4c ). Similarly, for parks, isolation does not affect species richness (t value = −1.677, p = 0.107, Figure S4a), Shannon index (t value = −1.866, p = 0.075, Figure S4b) or ant communities composition (F = 0.77, p = 0.69). Effect of park size As expected, larger parks harbour more species than smaller parks (F = 3.74, p = 0.039, R 2 = 0.19, Figure 3b ). The effect is significant between small and large parks (pairwise comparison: large-medium p = 0.21, medium-small p = 0.57, large-small p = 0.03). Large parks have on average 8 +/- 2.7 species, medium parks have 7.38 +/- 2.07 species and small ones have 7.17 +/- 2.14 species. Patch size has a similar but marginal influence on the Shannon index (F value = 2.83, p = 0.081, Figure 3e ). Patch size also affects the spatial turnover of ant communities (F value = 2.14, p = 0.008, R 2 = 16.3%, Figure 4b ). The beta diversity is composed of 100% of turnover between large and medium parks, 84% between medium and small, and 93% between large and small (Jaccard index large-medium = 0.222; medium-small = 0.278; large-small = 0.450) highlighting a dominant role of species sorting. Effect of urbanisation on species traits There is no effect of urbanisation on the total mass of a colony (t value = −0.76, p = 0.46, Figure 5a ). However, it is correlated with the social structure (z value = −2.26, p = 0.02, R2 = 0.15, Figure 5b ), with more monogynous species in urban habitats. It is also correlated with a higher meridional index, meaning that urban environments tend to select thermophilic ant species (t value = 2.89, p = 0.008, R2 = 0.21, Figure 5c ). Download figure Open in new tab Figure 5: Relation between urbanisation preference of the species and their traits: colony mass (a), Social structure (b), Meridional index as a proxy of thermophily (c). Urbanisation preference varies between zero for strictly forest species to one for strictly urban species. It is correlated to the social structure (z value = −2.26, p = 0.02, R 2 = 0.15) and thermophily (t value = 2.89, p = 0.008, R 2 = 0.21), but not to the mass of the colony (t value = −0.76, p = 0.46). Discussion Our study shows that urbanisation and the associated habitat fragmentation influence the litter ant communities at the metacommunity scale. Surprisingly, the ant biodiversity assessed by the total number of species per environment (gamma diversity) is stable, and the species richness per patch is slightly higher in parks than in forests. Communities differ markedly between the two environments, with a strong turnover in species compositions. At a local scale (alpha diversity), species richness and community composition are affected by park size and do not vary among forests. Finally, we highlight that the turnover in species composition between urban and forest patches is associated with systematic differences in species phenotypes. Urbanisation selects species that tend to be monogynous and thermophilic. A possible limitation of our study is that forests and urban parks were sampled in different years (2020 vs. 2021). While we acknowledge this issue, we however think that potential biases are here limited. A first reason stems from the fact that our method relies on community composition (presence/absence of species), not on species abundances. While abundances may fluctuate from a year to another, we do not expect such variations when looking at the presence/absence of long-lived species. Ant queens and colonies are indeed long-lived ( Hölldobler & Wilson, 1990 ). Ant queens live on average 10 years, up to 28 years, and colonies outlive queens in species where colonies have several queens or where queens can be replaced, e.g. by adoption of daughter queens ( Keller, 1998 ). The composition of ant species community is therefore relatively stable from one year to the other ( Donoso, 2017 ; Herbers, 2011 ). Focusing on species presence/absence, therefore reduces sensitivity to yearly fluctuations in colony size or activity. The clear differences observed between forests and parks align with ecological gradients (urbanisation, fragmentation, microclimate) rather than temporal effects. Notably, 2020 (forest sampling) was warmer in the Paris region than 2021 (park sampling, 2020: 14.3 °C, 2021: 12.9 °C. Source: opendata.paris.fr), yet we still detected thermophilisation in parks, which strengthens our inference. Finally, both habitats were sampled in the same season during no rain days, with sufficient effort to capture nearly all expected species (Fig. S3). Taken together, these points support that the observed differences reflect genuine ecological responses to urbanisation and fragmentation rather than temporal sampling artefacts. At a regional scale, comparing rural versus urban environments reveals no differences in the total number of species but a clear difference in community composition, with beta diversity dominated by its turnover component. This suggests that dispersal is low between the two environments, creating different communities adapted to each, and highlights species sorting as a strong regional-scale mechanism in this metacommunity. Indeed, if dispersal between forest and urban sites was high, we would expect either community composition to be similar (low beta diversity) or, if colonisation success differed between the environments, that the communities of one environment would be a subset of the communities of the other environment (high nestedness, ( Baselga, 2010 )). Differences in community composition between forests and parks can also reflect differences in fragmentation and shifts in ecological niches. In parks, wooded areas represent only a portion of the park and are often non-contiguous and surrounded by open habitats, allowing workers from these patches to forage in the sampled litter. In contrast, even the most fragmented forest patches were large enough to allow sampling away from the forest edge. Climate represents another niche axis: parks are warmer due to the urban heat island effect, attracting warm-adapted species, as confirmed by our results. Urbanisation also introduces various pollutants, which may select for particular species and traits. These niche changes, inherent to urban environments, likely contribute to the observed differences in community composition. Surprisingly, more species are present per urban park than per forest. Several explanations can be proposed. First, as mentioned above, species sorting dominates, suggesting restricted dispersal between forests and city parks, and parks may thus host more ant species than forests if they harbour a higher diversity of habitats and microenvironments. Although our sampling within parks targeted shaded litter patches under bushes or trees to mimic forest conditions, they were often surrounded by more open and trampled areas, potentially creating a range of microhabitats. This aspect, not fully quantifiable in our study due to a lack of direct microhabitat characterisation, could contribute to higher species richness in parks. However, if parks were richer simply because of a higher microhabitat heterogeneity, we would expect to find a high nestedness. Instead, our data reveal a near-complete turnover in species composition between forests and parks, suggesting that parks do not merely add species but rather replace forest species with species better adapted to urban conditions. Since warmer climates support more species, parks may receive a larger pool of colonisers, potentially increasing their species richness. This is confirmed with the trait analysis, where we have more warm-adapted species in the parks. Another possibility is that species richness in parks might be “inflated” by continuous colonisation processes such as soil inputs during park maintenance or transport of material within urban areas. These hypotheses would implicitly question our view of forests being good patches compared to parks. An alternative explanation would be that forests are (or have been until relatively recently) well-connected and form a metacommunity of their own. If so, high dispersal among forests could homogenise ant forest communities through mass effects, leading to a depressed lower alpha diversity reflecting partial homogenization ( Leibold et al., 2004 ; Mouquet & Loreau, 2003 ). In contrast, when dispersal is low, as may be the case between parks, competition occurs independently within each park resulting in different communities hence higher alpha and gamma diversities. This increases turnover between patches as observed in our data. This is likely given that parks were made for human leisure and differ more from one another than natural forests do. Our observations of local community compositions support the idea of strong mass effects among forests. Indeed, there is no difference in species richness and species composition among forests, even for the most fragmented ones. This suggests a high degree of dispersal between forest sites, which homogenises the communities. This is plausible because the matrix separating forests is less hostile than the one separating parks. Indeed, field hedges, roadsides and private gardens may connect forests by a network of trees and shrubs producing leaf litter and facilitating the dispersal of ants. Moreover, forests may be more similar to one another than parks. The former evolved from natural forests and have been managed largely for similar usages (e.g. wood production, leisure, hunting), whereas the latter are artificial and have to some extent been conceived to differ from one another to provide varied scenic environments. This could be confirmed by a more systematic study of biotic and abiotic environmental variations. If so, such homogeneity could limit the number of local niches, hence species sorting. In urban areas, on a park scale, species richness is generally lower in smaller parks, and community composition differs due to high species turnover. This pattern aligns with the classical species-area relationship, which links larger areas to higher diversity ( Preston, 1962 ), potentially reflecting intensified species sorting in smaller parks ( Leibold et al., 2004 ; Mouquet & Loreau, 2003 ). In smaller parks, competitive exclusion may be more pronounced, as only the most competitive species persist, while low dispersal limits the arrival of new species. This limited dispersal is exacerbated by the hostile urban matrix surrounding these parks, which restricts movement and colonisation, particularly in smaller parks less accessible through natural or human-mediated dispersal (e.g., fewer plants or materials introduced). Consequently, these combined niche and dispersal effects contribute to lower alpha diversity and higher turnover between smaller parks. These patterns are consistent with other studies, including other ant communities. Smaller parks in highly urbanized zones tend to harbour fewer ant species than those in peri-urban, less isolated areas ( Pacheco & Vasconcelos, 2007 ), while within cities, larger parks generally support greater alpha diversity ( Carpintero & Reyes-López, 2014 ). This trend extends to other ecosystems and taxa. For example, forests show similar patterns ( Leal et al., 2012 ). Other studies on isolated zooplankton communities in controlled environments also report reduced diversity due to limited dispersal ( Steiner & Asgari, 2022 ), mirroring the low mass effects in our urban ant metacommunity. Such findings are consistent with island biogeography theory, which suggests that isolated habitats, like urban parks or fragmented forests, host fewer species than more contiguous areas ( Ramalho et al., 2022 ). Beyond the turnover in species composition, we highlight that phenotypic traits are deterministically associated with community variations. Urban species are more often monogynous and more thermophilic. While the second trait was expected, given the heat island effect in cities, the first was not. Monogyny is somewhat linked to dispersal capacity ( Keller, 1991 ; Sundström et al., 2005 ). Monogynous colonies are often founded by winged queens that disperse alone by flying over long distances. In contrast, polygynous colonies are either founded by related non-flying queens that disperse on foot over short distances with workers or arise secondarily by non-flying queens that remain in their natal colony. Theoretically, habitat fragmentation can select for high dispersal (i.e. winged queens), for instance, when dispersal is traded off with competition capacity ( Finand, Monnin, et al., 2024 ; Tilman, 1994 ), when fragmentation leads to high kin competition in remaining patches, or to avoid inbreeding ( Cote et al., 2017 ; Gandon, 1999 ; Hamilton & May, 1977 ). Some empirical studies support these theoretical results, including in ants. For instance, at the intraspecific level, colonisation ability can be in a trade-off with competition in an ant species ( Finand, Loeuille, et al., 2024 ). This selection for high dispersal in a fragmented context could explain, as a byproduct, that city parks harbour more monogynous species because polygynous species disperse less. However, some studies have shown opposite results, with habitat fragmentation in cities selecting for lower dispersal ( Cheptou et al., 2008 ; Finand et al., 2023 ). Indeed, while habitat fragmentation usually favours high dispersal, it may have the opposite effect under some circumstances (e.g. highly hostile matrix increasing dispersal costs, Bonte et al. 2012 ; Hastings 1983 ). A second explanation is that monogynous species can be considered as pursuing spatial bet-hedging because flying queens can access and distribute the risks on more patches than walking queens. Indeed, a colony producing flying queens can attempt to establish new colonies on many patches. In contrast, a colony producing apterous queens produces a few only, as each requires the allocation of workers, and they have restricted dispersal, hence it has few establishment attempts. A bet-hedging strategy is selected under temporally variable environments ( Slatkin, 1974 ). City patches may be more variable, due to many disturbances associated with human activities, and parks surrounded by unfavourable environments. Having winged queens will increase the probability of having offspring in a more favourable environment. A third hypothesis is that urban parks are relatively new, particularly regarding ants that have a longer generation time than most other insects. The first species colonising a new patch are the most dispersive ones, which in ants have winged queens and thus are monogynous. Dispersal limitation also likely lasts in this highly fragmented context. Forests being much older, the less dispersive but more competitive species had the time to colonise them and progressively exclude better colonisers. All these hypotheses are not exclusive. Other taxa have shown this shift in community composition with more dispersive species in the city, like carabid beetles ( Piano et al., 2017 ). The high abundance of thermophilic ants in urban areas aligns with numerous studies indicating that cities, such as Paris, are subject to the Urban Heat Island effect ( Lemonsu et al., 2015 ). Heat-susceptible species may disappear, and thermophilic species may dominate communities in cities. Parr and Bishop (2022) suggest that temperate ant species may be less vulnerable to climate warming than tropical species, and in some cases might even benefit from rising temperatures. However, our findings—showing a near-complete replacement of forest ant communities by species with more southern affinities in urban parks—do not fully support this idea. On the contrary, although temperature is not the only factor differing between urban and rural forests, our results suggest that our temperate ant communities living in forests may be negatively impacted by rising temperatures. The prevalence of thermophilic species in urban areas has been documented not only in ants but also in other groups, such as carabid beetles ( Piano et al., 2017 ) and moths ( Franzén et al., 2020 ). More broadly, the process of thermophilisation, where communities shift toward species that prefer warmer conditions due to rising temperatures, has also been observed in non-urban ecosystems, including fish ( Daufresne et al., 2004 ) and plants ( Feeley et al., 2020 ). This result is particularly relevant in the context of climate change, as temperatures are expected to rise in the coming decades, leading to significant ecological consequences ( McCarty, 2001 ). Cities offer a valuable natural laboratory for studying the impacts of climate change, with urban communities serving as ideal experimental units to investigate these effects. Furthermore, our findings suggest that climate change could act as a critical filter for community composition, potentially leading to the decline of less thermophilic species. While urban areas might serve as reservoirs for future biodiversity under changing climates, they will likely become increasingly inhospitable for many species as climate change progresses. Acknowledgements We would like to thank the City of Paris and the Natural History Museum of Paris for allowing us to sample the parks of the city. We also thank Xavier Espadaler and Bernard Kaufmann for their help in the identification of some species. Footnotes Conflict of interest: The authors declare no conflict of interest. Authorship statement: Basile Finand, Thibaud Monnin, and Nicolas Loeuille conceived the study. Basile Finand, Thibaud Monnin, Nicolas Loeuille, Léa Darmedru, Joséphine Ledamoisel, Céline Bocquet, Angélique Bultelle, Pierre Fédérici, and Solène Gouffault collected the data. Basile Finand analysed the data. Basile Finand, Thibaud Monnin, and Nicolas Loeuille wrote the paper. All the other co-authors agreed on the last version of the paper. 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Share Urbanisation and habitat fragmentation favour thermophilic and monogynous ant species Basile Finand , Thibaud Monnin , Céline Bocquet , Angélique Bultelle , Pierre Fédérici , Léa Darmedru , Solène Gouffault , Joséphine Ledamoisel , Nicolas Loeuille bioRxiv 2025.02.25.640022; doi: https://doi.org/10.1101/2025.02.25.640022 Share This Article: Copy Citation Tools Urbanisation and habitat fragmentation favour thermophilic and monogynous ant species Basile Finand , Thibaud Monnin , Céline Bocquet , Angélique Bultelle , Pierre Fédérici , Léa Darmedru , Solène Gouffault , Joséphine Ledamoisel , Nicolas Loeuille bioRxiv 2025.02.25.640022; doi: https://doi.org/10.1101/2025.02.25.640022 Citation Manager Formats BibTeX Bookends EasyBib EndNote (tagged) EndNote 8 (xml) Medlars Mendeley Papers RefWorks Tagged Ref Manager RIS Zotero Tweet Widget Facebook Like Google Plus One Subject Area Ecology Subject Areas All Articles Animal Behavior and Cognition (7635) Biochemistry (17691) Bioengineering (13892) Bioinformatics (41937) Biophysics (21452) Cancer Biology (18588) Cell Biology (25504) Clinical Trials (138) Developmental Biology (13378) Ecology (19899) Epidemiology (2067) Evolutionary Biology (24320) Genetics (15609) Genomics (22506) Immunology (17736) Microbiology (40394) Molecular Biology (17181) Neuroscience (88605) Paleontology (666) Pathology (2832) Pharmacology and Toxicology (4824) Physiology (7641) Plant Biology (15156) Scientific Communication and Education (2045) Synthetic Biology (4294) Systems Biology (9825) Zoology (2271)