# Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area

Giovanny Destour, Rumsais Blatrix, Aurélien Caries, Juliette Genevet, Bernard Kaufmann, Baptiste Lecoq, Annick Lucas, Killian Verlingue, Alan Vergnes, Marion Javal

<figure><img src="image_1.png" /></figure>

To cite this version:

Giovanny Destour, Rumsais Blatrix, Aurélien Caries, Juliette Genevet, Bernard Kaufmann, et al. Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area. Ecological Entomology, In press, 10.1111/een.13339. hal-04597120

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Submitted on 1 Jun 2024

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# Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area

Giovanny Destour, Rumsaïs Blatrix, Aurélien Caries, Juliette Genevet, Bernard Kaufmann, Baptiste Lecoq, Annick Lucas, Killian Verlingue, Alan Vergnes, Marion Javal

<figure><img src="image_1.png" /></figure>

## To cite this version:

Giovanny Destour, Rumsaïs Blatrix, Aurélien Caries, Juliette Genevet, Bernard Kaufmann, et al. Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area. Ecological Entomology, 2024, 49, pp.673-681. 10.1111/een.13339. hal-04597120v2

HAL Id: hal-04597120

https://hal.inrae.fr/hal-04597120v2

Submitted on 23 Sep 2024

HAL is a multi-disciplinary open access archive for the deposit and dissemination of scientific research documents, whether they are published or not. The documents may come from teaching and research institutions in France or abroad, or from public or private research centers.

L'archive ouverte pluridisciplinaire HAL, est destinée au dépôt et à la diffusion de documents scientifiques de niveau recherche, publiés ou non, émanant des établissements d'enseignement et de recherche français ou étrangers, des laboratoires publics ou privés.

<figure><img src="image_2.png" /></figure>

Distributed under a Creative Commons Attribution 4.0 International License
ORIGINAL ARTICLE

Ecological
Entomology

<figure><img src="image_3.png" /></figure>

# Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area

Giovanny Destour<fnref n="1" /> ID

Rumsais Blatrix<fnref n="1" /> ID

Aurélien Caries<fnref n="1" /> ID

Juliette Genevet<fnref n="1" />

Bernard Kaufmann<fnref n="2" /> ID

Baptiste Lecoq<fnref n="1" /> ID

Annick Lucas<fnref n="1" />

Killian Verlingue<fnref n="1" />

Alan Vergnes<fnref n="3" /> ID

Marion Javal<fnref n="1" />,<fnref n="4" /> ID

**Correspondence**

Marion Javal, CEFE, Université de Montpellier, CNRS, EPHE, IRD, 34293 Montpellier, France.

Email: marion.javal@ird.fr

**Funding information**

Montpellier Méditerranée Metropolis and the City of Montpellier through the Urban ecology agreement with CEFE.

Associate Editor: Tali Reiner Brodetzki

**Abstract**

1. Invasive species often have harmful effects on human activities and biodiversity. The *Tapinoma nigerrimum* complex includes three species that exhibit invasive traits and the ability to form supercolonies, including *Tapinoma magnum* and *Tapinoma darioi*. While these species are native to the Mediterranean basin, they have recently been reported as invasive in more northern regions of Europe. However, *T. magnum* and *T. darioi* show different invasion patterns in Europe, suggesting differences in thermal preferences between these species. We aimed to determine whether there was any difference in activity pattern depending on temperature or season between the two *Tapinoma* species and between *Tapinoma* and co-occurring native Mediterranean ants.

2. We measured ambient temperature and ant foraging activity around the nest entrances of *T. darioi* and *T. magnum* in urbanised sites around Montpellier, southern France, over almost 1 year.

3. We did not detect differences in activity between the two species according to temperature, but *T. magnum* was more active than *T. darioi* in winter. Compared to the co-occurring native species, the two *Tapinoma* species were more active at low temperatures and earlier in the year.

4. These results suggest that the difference in invasion success between *T. magnum* and *T. darioi* may not be explained by temperature preference. The difference in activity between native Mediterranean species and invasive *Tapinoma* suggests a coexistence through niche partitioning. In addition, our results suggest that control measures against invasive *Tapinoma* species would be more effective if applied at the end of winter.

**KEYWORDS**

ants, biological invasions, coexistence, establishment, exploitation competition, foraging

This is an open access article under the terms of the [Creative Commons Attribution](https://creativecommons.org/licenses/by) License, which permits use, distribution and reproduction in any medium, provided the original work is properly cited.

© 2024 The Authors. *Ecological Entomology* published by John Wiley & Sons Ltd on behalf of Royal Entomological Society.

<figure><img src="image_4.png" /></figure>

<footnote n="1">¹CEFE, Université de Montpellier, CNRS, EPHE, IRD, Montpellier, France</footnote>

<footnote n="2">²Université Claude Bernard Lyon 1, LEHNA UMR 5023, CNRS, ENTPE, Villeurbanne, France</footnote>

<footnote n="3">³CEFE, Université Paul Valéry Montpellier 3, Université de Montpellier, CNRS, EPHE, IRD, Montpellier, France</footnote>

<footnote n="4">⁴CBGP, IRD, CIRAD, INRAE, Institut Agro Montpellier, Université de Montpellier, Montpellier, France</footnote>
ORIGINAL ARTICLE

Ecological
Entomology

<figure><img src="image_3.png" /></figure>

# Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area

Giovanny Destour<fnref n="1" /> ID

Rumsais Blatrix<fnref n="1" /> ID

Aurélien Caries<fnref n="1" /> ID

Juliette Genevet<fnref n="1" />

Bernard Kaufmann<fnref n="2" /> ID

Baptiste Lecoq<fnref n="1" /> ID

Annick Lucas<fnref n="1" />

Killian Verlingue<fnref n="1" />

Alan Vergnes<fnref n="3" /> ID

Marion Javal<fnref n="1" />,<fnref n="4" /> ID

**Correspondence**

Marion Javal, CEFE, Université de Montpellier, CNRS, EPHE, IRD, 34293 Montpellier, France.
Email: marion.javal@ird.fr

**Funding information**

Montpellier Méditerranée Metropolis and the City of Montpellier through the Urban ecology agreement with CEFE.

Associate Editor: Tali Reiner Brodetzki

**Abstract**

1. Invasive species often have harmful effects on human activities and biodiversity. The *Tapinoma nigerrimum* complex includes three species that exhibit invasive traits and the ability to form supercolonies, including *Tapinoma magnum* and *Tapinoma darioi*. While these species are native to the Mediterranean basin, they have recently been reported as invasive in more northern regions of Europe. However, *T. magnum* and *T. darioi* show different invasion patterns in Europe, suggesting differences in thermal preferences between these species. We aimed to determine whether there was any difference in activity pattern depending on temperature or season between the two *Tapinoma* species and between *Tapinoma* and co-occurring native Mediterranean ants.

2. We measured ambient temperature and ant foraging activity around the nest entrances of *T. darioi* and *T. magnum* in urbanised sites around Montpellier, southern France, over almost 1 year.

3. We did not detect differences in activity between the two species according to temperature, but *T. magnum* was more active than *T. darioi* in winter. Compared to the co-occurring native species, the two *Tapinoma* species were more active at low temperatures and earlier in the year.

4. These results suggest that the difference in invasion success between *T. magnum* and *T. darioi* may not be explained by temperature preference. The difference in activity between native Mediterranean species and invasive *Tapinoma* suggests a coexistence through niche partitioning. In addition, our results suggest that control measures against invasive *Tapinoma* species would be more effective if applied at the end of winter.

**KEYWORDS**

ants, biological invasions, coexistence, establishment, exploitation competition, foraging

<figure><img src="image_4.png" /></figure>

<footnote n="1">¹CEFE, Université de Montpellier, CNRS, EPHE, IRD, Montpellier, France</footnote>

<footnote n="2">²Université Claude Bernard Lyon 1, LEHNA UMR 5023, CNRS, ENTPE, Villeurbanne, France</footnote>

<footnote n="3">³CEFE, Université Paul Valéry Montpellier 3, Université de Montpellier, CNRS, EPHE, IRD, Montpellier, France</footnote>

<footnote n="4">⁴CBGP, IRD, CIRAD, INRAE, Institut Agro Montpellier, Université de Montpellier, Montpellier, France</footnote>

<footnote>This is an open access article under the terms of the [Creative Commons Attribution](https://creativecommons.org/licenses/by) License, which permits use, distribution and reproduction in any medium, provided the original work is properly cited.

© 2024 The Authors. *Ecological Entomology* published by John Wiley & Sons Ltd on behalf of Royal Entomological Society.</footnote>
INTRODUCTION

Invasive species, capable of proliferating rapidly and colonising new environments outside of their native range, are a threat to biodiversity, economy and health worldwide (Diagne et al., 2021; McGeoch et al., 2010; Simberloff et al., 2013). The spread of these species is mainly favoured by global change and the increase in trade and there is no sign that this trend is slowing down on a global scale, particularly in the case of insects (Seebens et al., 2017). However, not all transported species necessarily become invasive, and their invasion success can be compromised at different stages of the process (Williamson & Fitter, 1996). The success of the invasion depends on intrinsic factors (e.g., propagule pressure, or life history traits of the introduced species), as well as the biotic and abiotic conditions of their new environment (Young et al., 2017). The climatic parameters of the invaded area are among the most important abiotic determinants of the establishment of invasive species (Pyšek et al., 2020) and may or may not allow the invasion to succeed by affecting the survival, growth or reproduction of these species (Dijkstra et al., 2017; Lord, 2017; Shi et al., 2010). Sometimes, the introduced invasive species will also be able to establish in the new environment by shifting its ecological niche because of divergent environmental conditions (Bates & Bertelsmeier, 2021; Broennimann et al., 2007; Tang et al., 2021). This phenomenon involves shifts of the introduced species biology (e.g., modification of phenotypic traits, accelerated or delayed phenology, changes in tolerance to environmental parameters).

Once established, these species can modify interspecific interac-
tions by outcompeting native species and disrupting the invaded eco-
system (Didham et al., 2005). However, native species can persist in
the invaded area if biological modifications result in minimising the
impact of the invasive species on their fitness (Berthon, 2015). To
avoid or reduce competition with the invasive species, for example,
native species can change their behaviour (Ruland & Jeschke, 2020).
Microclimatic and temporal heterogeneity play a role in interspecific
interactions between native and invasive species. Depending on envi-
ronmental or seasonal factors, native species can adjust activity pat-
terns or habitat selection to avoid competition or outcompete
invasive species (Gippet et al., 2022; Melbourne et al., 2007; Perales
et al., 2021; Valladares et al., 2015). Environmental factors, such as
temperature and seasonality, influence foraging and can also allow the
coexistence of competing native and invasive species if the prefer-
ences of each are different (Albrecht & Gotelli, 2001; Roeder
et al., 2018).

To date, over 200 species of ants (out of circa 14,000 described
species) are considered as invasive, including 5 species belonging to
the top 100 most invasive species worldwide (Bertelsmeier
et al., 2018; Lowe et al., 2000). Invasive ants generally share some
behavioural characteristics such as the capacity to form supercolonies,
which consists of an interconnection of numerous nests (polydomy)
housing large numbers of queens (polygyny) and workers
(Helanterä, 2022; Moffett, 2012). Some of them are considered as
tramp species because they are accidentally introduced by human
activities and are harmful to ecosystems (Passera, 1994). Areas

invaded by invasive ant species are therefore often overpopulated by
these species to the detriment of other organisms (Gan et al., 2022;
Mikissa et al., 2013; Wetterer & Moore, 2005). Invasive ant species
are particularly good at locating and exploiting resources, and when
foraging they tend to monopolise food sources (Davidson, 1998).
These species are often highly competitive and exclude local species
through direct aggression or more efficient foraging (Holway
et al., 2002).

The *Tapinoma nigerrimum* complex comprises four ant species, *T. darioi*, *T. ibericum*, *T. magnum* and *T. nigerrimum*, all of which originate from the Mediterranean basin. These species are very similar morphologically and were considered as a single species until recently (Seifert et al., 2017). This complex includes three polydomous species (*T. magnum*, *T. darioi* and *T. ibericum*) which can form supercolonies (Seifert et al., 2017). Among the few studies carried out on the ants of this complex, some show the strong competitiveness of *T. magnum* and its negative impact on native ants (Blight et al., 2010; Gippet et al., 2022).

In this study, we focus on *T. magnum* and *T. darioi*, two potential
tramp species prevalent in anthropised environments, which appear
to be spreading through human activities (Centanni et al., 2022;
Lenoir et al., 2023; Seifert et al., 2017). Although they appear very
similar morphologically and behaviourally, they differ in their sup-
posed native range and invasion pattern. The supposed native range
of *T. magnum* covers North Africa, Italy and the Mediterranean coast
of southeastern France. *Tapinoma darioi* has a supposed native
range extending from the Mediterranean coast of southwestern
France to the northeast of Spain (Centanni et al., 2022; Gouraud &
Kaufmann, 2022; Seifert et al., 2017). The two species have been
introduced in regions north of the Mediterranean, *T. magnum* having
been detected in Belgium, Germany, Switzerland and the
Netherlands, *T. darioi* in Switzerland (Dekoninck et al., 2015; Lenoir
et al., 2023; Seifert et al., 2017). These species are also invasive in
several French localities, even in Mediterranean areas and are in
constant expansion (Centanni et al., 2022; Gouraud &
Kaufmann, 2022; Lenoir et al., 2023). To date, reports of introduc-
tions of these species appear to be limited to urban areas. In addi-
tion, Centanni et al. (2022) showed that, in the region of
Montpellier, southern France, *T. darioi* occurred more frequently in
environments with an intermediate level of urbanisation. Thus, there
may be few conservation issues related to the introduction of these
species. Nevertheless, it does represent a threat to human activities
(e.g., harvest loss in market gardening, disruption of the use of green
spaces in cities), as evidenced by the growing number of articles in
the local press in mainland France and Corsica. In addition, although
*T. darioi* is probably native to the dune environment along the Medi-
terranean French coast, it thrives in this fragile ecosystem (Centanni
et al., 2022), and neither the effect of the introduction of new
genetic groups on the dynamics of the species nor the potential
consequences on local species in their native range are known yet.
In this context, understanding the factors underlying the expansion
of *T. darioi* and *T. magnum* would be useful for management
strategies.
# INTRODUCTION

Invasive species, capable of proliferating rapidly and colonising new environments outside of their native range, are a threat to biodiversity, economy and health worldwide (Diagne et al., 2021; McGeoch et al., 2010; Simberloff et al., 2013). The spread of these species is mainly favoured by global change and the increase in trade and there is no sign that this trend is slowing down on a global scale, particularly in the case of insects (Seebens et al., 2017). However, not all transported species necessarily become invasive, and their invasion success can be compromised at different stages of the process (Williamson & Fitter, 1996). The success of the invasion depends on intrinsic factors (e.g., propagule pressure, or life history traits of the introduced species), as well as the biotic and abiotic conditions of their new environment (Young et al., 2017). The climatic parameters of the invaded area are among the most important abiotic determinants of the establishment of invasive species (Pyšek et al., 2020) and may or may not allow the invasion to succeed by affecting the survival, growth or reproduction of these species (Dijkstra et al., 2017; Lord, 2017; Shi et al., 2010). Sometimes, the introduced invasive species will also be able to establish in the new environment by shifting its ecological niche because of divergent environmental conditions (Bates & Bertelsmeier, 2021; Broennimann et al., 2007; Tang et al., 2021). This phenomenon involves shifts of the introduced species biology (e.g., modification of phenotypic traits, accelerated or delayed phenology, changes in tolerance to environmental parameters).

Once established, these species can modify interspecific interactions by outcompeting native species and disrupting the invaded ecosystem (Didham et al., 2005). However, native species can persist in the invaded area if biological modifications result in minimising the impact of the invasive species on their fitness (Berthon, 2015). To avoid or reduce competition with the invasive species, for example, native species can change their behaviour (Ruland & Jeschke, 2020). Microclimatic and temporal heterogeneity play a role in interspecific interactions between native and invasive species. Depending on environmental or seasonal factors, native species can adjust activity patterns or habitat selection to avoid competition or outcompete invasive species (Gippet et al., 2022; Melbourne et al., 2007; Perales et al., 2021; Valladares et al., 2015). Environmental factors, such as temperature and seasonality, influence foraging and can also allow the coexistence of competing native and invasive species if the preferences of each are different (Albrecht & Gotelli, 2001; Roeder et al., 2018).

To date, over 200 species of ants (out of circa 14,000 described species) are considered as invasive, including 5 species belonging to the top 100 most invasive species worldwide (Bertelsmeier et al., 2018; Lowe et al., 2000). Invasive ants generally share some behavioural characteristics such as the capacity to form supercolonies, which consists of an interconnection of numerous nests (polydomy) housing large numbers of queens (polygyny) and workers (Helanterä, 2022; Moffett, 2012). Some of them are considered as tramp species because they are accidentally introduced by human activities and are harmful to ecosystems (Passera, 1994).

Areas invaded by invasive ant species are therefore often overpopulated by these species to the detriment of other organisms (Gan et al., 2022; Mikissa et al., 2013; Wetterer & Moore, 2005). Invasive ant species are particularly good at locating and exploiting resources, and when foraging they tend to monopolise food sources (Davidson, 1998). These species are often highly competitive and exclude local species through direct aggression or more efficient foraging (Holway et al., 2002).

The *Tapinoma nigerrimum* complex comprises four ant species, *T. darioi*, *T. ibericum*, *T. magnum* and *T. nigerrimum*, all of which originate from the Mediterranean basin. These species are very similar morphologically and were considered as a single species until recently (Seifert et al., 2017). This complex includes three polydomous species (*T. magnum*, *T. darioi* and *T. ibericum*) which can form supercolonies (Seifert et al., 2017). Among the few studies carried out on the ants of this complex, some show the strong competitiveness of *T. magnum* and its negative impact on native ants (Blight et al., 2010; Gippet et al., 2022).

In this study, we focus on *T. magnum* and *T. darioi*, two potential tramp species prevalent in anthropised environments, which appear to be spreading through human activities (Centanni et al., 2022; Lenoir et al., 2023; Seifert et al., 2017). Although they appear very similar morphologically and behaviourally, they differ in their supposed native range and invasion pattern. The supposed native range of *T. magnum* covers North Africa, Italy and the Mediterranean coast of southeastern France. *Tapinoma darioi* has a supposed native range extending from the Mediterranean coast of southwestern France to the northeast of Spain (Centanni et al., 2022; Gouraud & Kaufmann, 2022; Seifert et al., 2017). The two species have been introduced in regions north of the Mediterranean, *T. magnum* having been detected in Belgium, Germany, Switzerland and the Netherlands, *T. darioi* in Switzerland (Dekoninck et al., 2015; Lenoir et al., 2023; Seifert et al., 2017). These species are also invasive in several French localities, even in Mediterranean areas and are in constant expansion (Centanni et al., 2022; Gouraud & Kaufmann, 2022; Lenoir et al., 2023). To date, reports of introductions of these species appear to be limited to urban areas. In addition, Centanni et al. (2022) showed that, in the region of Montpellier, southern France, *T. darioi* occurred more frequently in environments with an intermediate level of urbanisation. Thus, there may be few conservation issues related to the introduction of these species. Nevertheless, it does represent a threat to human activities (e.g., harvest loss in market gardening, disruption of the use of green spaces in cities), as evidenced by the growing number of articles in the local press in mainland France and Corsica. In addition, although *T. darioi* is probably native to the dune environment along the Mediterranean French coast, it thrives in this fragile ecosystem (Centanni et al., 2022), and neither the effect of the introduction of new genetic groups on the dynamics of the species nor the potential consequences on local species in their native range are known yet. In this context, understanding the factors underlying the expansion of *T. darioi* and *T. magnum* would be useful for management strategies.
<figure><img src="image_5.png" /><figcaption>**FIGURE 1** Locations of sampling sites in the department of Hérault are marked in red. The green triangles represent *Tapinoma magnum* colonies and the orange discs *Tapinoma darioi* colonies monitored in this study. The map was generated using QGIS 3.28.3.</figcaption></figure>

The two invasive species are present in different proportions in the various areas invaded. For example, Centanni et al. (2022) showed a higher occurrence of *T. darioi* compared with *T. magnum* in southern France. In contrast, *T. magnum* is more common in the northern and colder regions of Europe (Lenoir et al., 2023; Seifert et al., 2017). As ants are ectothermic organisms (Angilletta, 2009; Ruano et al., 2000), temperature is one of the main determinants of their activity, including foraging (Pol & de Casenave, 2004). Given the predominance of *T. magnum* compared to *T. darioi* in the northern regions of Europe and the high acclimatisation capacity of *T. magnum* (Bujan et al., 2021), we hypothesise a difference in thermal preference between the two species. Specifically, we postulate that *T. magnum* is active at lower temperatures than *T. darioi*. If so, this could explain the higher proportion of *T. magnum* in northern localities of Western Europe. Moreover, these invasive species could have an impact on the native ants, and we sought to take a step forward in understanding the interactions of these species with the native species present. Therefore, we aimed to determine whether there was any difference in activity pattern between the two invasive *Tapinoma* species and co-occurring native Mediterranean ants, which could allow the native species to persist in the invaded areas. To answer these questions, we compared foraging activity between invasive *Tapinoma* species at the nest entrance during regular monitoring over almost a year, covering a wide range of temperatures. We also compared the foraging activity of the two invasive *Tapinoma* with that of the other ants present around the *Tapinoma* nest entrances.

# MATERIALS AND METHODS

## Study area and experimental design

Our study is located in Southern France, in the Hérault department, around the city of Montpellier. The climate is semi-arid Mediterranean. The area in and around the city of Montpellier is highly

<figure><img src="image_6.png" /><figcaption>**FIGURE 2** Experimental design of the activity measurement protocol. The baits are represented by white squares placed around the central point of activity of the *Tapinoma* colony represented by a red disc.</figcaption></figure>

urbanised and densely populated, with more than 5000 inhabitants per km² (INSEE, 2022).

For activity measurements, three colonies of each species were selected around the city of Montpellier, based on Centanni et al. (2022), who identified colonies at the species level using microsatellite data. *Tapinoma magnum* colonies being rarer around Montpellier (Centanni et al., 2022), three colonies of this species were first selected (*T. magnum* colonies: 43°39′13.9" N 3°55′11.5" E, 43°35′17.8" N 3°56′38.7" E, 43°32′46.0" N 4°07′27.5" E) (Figure 1). Then, to use *T. darioi* colonies exposed to a range of climatic conditions comparable to that of the *T. magnum* colonies, we selected one colony of *T. darioi* close (400–2000 m) to each *T. magnum* colony (*T. darioi* colonies: 43°39′11.6" N 3°54′31.6" E, 43°35′25.2" N 3°56′53.3" E, 43°33′14.3" N 4°06′31.0" E) (Figure 1). The nests were located in green spaces in an urbanised environment: flowerbeds, wasteland and beachside (Table S1).

For each activity measurement, 18 baits were laid out in two parallel lines of 9 baits spaced 1 metre apart, each line being centred on a central point (main entrance of the *Tapinoma* nest), and separated by a distance of 40 cm (Figure 2). Each bait consisted of a strip of paper (11 × 4 cm), on which a drop of honey, tuna rillettes and peanut butter were placed, which are baits commonly used to attract ants (Brinkman et al., 2001; Rahardjo et al., 2023; Wyatt, 2004). After 30 min, the baits with ants were photographed in order to count and identify the ants present (species of the *Tapinoma* colonies were identified in a previous study, see above, and native species were identified on photos). Between 22 and 30 measurement sessions were carried out depending on the colony (147 field sessions in total). The measurement sessions were carried out between autumn 2021 and summer 2022, mostly between January and April (see Figure 3b for the temporal distribution of the sampling).

Soil temperature was measured at the beginning of each foraging activity measurement session at the central point and 50 cm on both sides of this point using a probe thermometer (CookConcept, ±0.5°C) inserted in the ground at a depth of 10 cm, and averaged over the three measurements for each session (Table S2).

## Statistical analysis

To test foraging activity of ants throughout the year, activity measurements were processed using generalised additive mixed models
<figure><img src="image_5.png" /><figcaption>**FIGURE 1** Locations of sampling sites in the department of Hérault are marked in red. The green triangles represent *Tapinoma magnum* colonies and the orange discs *Tapinoma darioi* colonies monitored in this study. The map was generated using QGIS 3.28.3.</figcaption></figure>

The two invasive species are present in different proportions in the various areas invaded. For example, Centanni et al. (2022) showed a higher occurrence of *T. darioi* compared with *T. magnum* in southern France. In contrast, *T. magnum* is more common in the northern and colder regions of Europe (Lenoir et al., 2023; Seifert et al., 2017). As ants are ectothermic organisms (Angilletta, 2009; Ruano et al., 2000), temperature is one of the main determinants of their activity, including foraging (Pol & de Casenave, 2004). Given the predominance of *T. magnum* compared to *T. darioi* in the northern regions of Europe and the high acclimatisation capacity of *T. magnum* (Bujan et al., 2021), we hypothesise a difference in thermal preference between the two species. Specifically, we postulate that *T. magnum* is active at lower temperatures than *T. darioi*. If so, this could explain the higher proportion of *T. magnum* in northern localities of Western Europe. Moreover, these invasive species could have an impact on the native ants, and we sought to take a step forward in understanding the interactions of these species with the native species present. Therefore, we aimed to determine whether there was any difference in activity pattern between the two invasive *Tapinoma* species and co-occurring native Mediterranean ants, which could allow the native species to persist in the invaded areas. To answer these questions, we compared foraging activity between invasive *Tapinoma* species at the nest entrance during regular monitoring over almost a year, covering a wide range of temperatures. We also compared the foraging activity of the two invasive *Tapinoma* with that of the other ants present around the *Tapinoma* nest entrances.

# MATERIALS AND METHODS

## Study area and experimental design

Our study is located in Southern France, in the Hérault department, around the city of Montpellier. The climate is semi-arid Mediterranean. The area in and around the city of Montpellier is highly

<figure><img src="image_6.png" /><figcaption>**FIGURE 2** Experimental design of the activity measurement protocol. The baits are represented by white squares placed around the central point of activity of the *Tapinoma* colony represented by a red disc.</figcaption></figure>

urbanised and densely populated, with more than 5000 inhabitants per km² (INSEE, 2022).

For activity measurements, three colonies of each species were selected around the city of Montpellier, based on Centanni et al. (2022), who identified colonies at the species level using microsatellite data. *Tapinoma magnum* colonies being rarer around Montpellier (Centanni et al., 2022), three colonies of this species were first selected (*T. magnum* colonies: 43°39′13.9" N 3°55′11.5" E, 43°35′17.8" N 3°56′38.7" E, 43°32′46.0" N 4°07′27.5" E) (Figure 1). Then, to use *T. darioi* colonies exposed to a range of climatic conditions comparable to that of the *T. magnum* colonies, we selected one colony of *T. darioi* close (400–2000 m) to each *T. magnum* colony (*T. darioi* colonies: 43°39′11.6" N 3°54′31.6" E, 43°35′25.2" N 3°56′53.3" E, 43°33′14.3" N 4°06′31.0" E) (Figure 1). The nests were located in green spaces in an urbanised environment: flowerbeds, wasteland and beachside (Table S1).

For each activity measurement, 18 baits were laid out in two parallel lines of 9 baits spaced 1 metre apart, each line being centred on a central point (main entrance of the *Tapinoma* nest), and separated by a distance of 40 cm (Figure 2). Each bait consisted of a strip of paper (11 × 4 cm), on which a drop of honey, tuna rillettes and peanut butter were placed, which are baits commonly used to attract ants (Brinkman et al., 2001; Rahardjo et al., 2023; Wyatt, 2004). After 30 min, the baits with ants were photographed in order to count and identify the ants present (species of the *Tapinoma* colonies were identified in a previous study, see above, and native species were identified on photos). Between 22 and 30 measurement sessions were carried out depending on the colony (147 field sessions in total). The measurement sessions were carried out between autumn 2021 and summer 2022, mostly between January and April (see Figure 3b for the temporal distribution of the sampling).

Soil temperature was measured at the beginning of each foraging activity measurement session at the central point and 50 cm on both sides of this point using a probe thermometer (CookConcept, ±0.5°C) inserted in the ground at a depth of 10 cm, and averaged over the three measurements for each session (Table S2).

## Statistical analysis

To test foraging activity of ants throughout the year, activity measurements were processed using generalised additive mixed models
<figure><img src="image_7.png" /><figcaption>**FIGURE 3** Generalised additive mixed model predictions of *Tapinoma magnum* workers (green dotted line, *n* = 76 measurement sessions) and *Tapinoma darioi* workers (orange solid line, *n* = 71 measurement sessions) average abundance scores, with 95% confidence intervals, according to (a) temperature and (b) day of the year (black vertical bars above the x-axis represent sampling days and coloured disk the start of a season). The abundance score of ants is classified into five categories corresponding to the average number of individuals on the 18 baits (0 = 0 worker; 1 = 1 worker; 2 = 2-5 workers; 3 = 6-20 workers; 4 = 21-100 workers; 5 = more than 100 workers). The grey area represents the time period when the activities of *T. magnum* and *T. darioi* differed significantly.</figcaption></figure>

(GAMMs) (Wood, 2006). This semi-parametric model is a flexible regression method for estimating linear and non-linear trends. GAMMs are often used to predict smooth functional relationships between the response variable and the explanatory variables, with the advantage that there are no predefined shape for describing functional relationships (Pedersen et al., 2019; Wood, 2017). Additionally, the foraging activity measurements followed an unevenly spaced time series because the measurement sessions were irregularly spaced in time, which made it difficult to apply conventional time series methods (Simpson, 2018). According to the methods of Andersen (1992), an abundance score (0 = 0 worker; 1 = 1 worker; 2 = 2-5 workers; 3 = 6-20 workers; 4 = 21-100 workers; 5 = more than 100 workers) was calculated for each bait occupied by the ant species considered. Then the average abundance score on the 18 baits was calculated and used as the foraging activity measurement because using the number of individuals recorded yielded overdispersed residuals in the models.

First, independent GAMMs were parameterised for *T. magnum* and for *T. darioi*, respectively, including as response variable the average abundance score and as explanatory variable the mean temperature with the smooth thin plate regression splines term and the site as random effect. Thin plate regression splines can be used to model complex non-linear relationships between variables, smoothing the data efficiently and flexibly without having to specify the location of the nodes (Wood, 2003). Two new GAMMs (one for each species) were parameterised with the same variables, but replacing the mean temperature by the day of the year as explanatory variable with the smooth cyclic cubic spline term. This smoothing term forces the ends of the spline to be equal to each other, allowing continuity between the first and last day of the year (Wood, 2017). In a second phase, the same GAMMs were carried out, considering the activities of both

invasive *Tapinoma* species together to infer their general activity pattern and compare it with that of native species.

This approach provides a reliable estimate by using 95% confidence limits around the modelled responses which we used to determine the differences in activity between ants in different categories (i.e., *T. magnum* vs. *T. darioi* and *Tapinoma* vs. native ants). For all the models, the distribution of the specified errors was quasi-Poisson. Model parameters were chosen by restricted maximum likelihood (REML) selection. All the GAMMs were produced using the ‘gam’ function in the ‘mgcv’ package (Wood, 2011) using R 4.2.1 software (R Core Team, 2022) and R studio (RStudio Core Team, 2020).

# RESULTS

## Species detected on baits

Overall, nine genera other than *Tapinoma* were recorded on the baits: *Camponotus*, *Crematogaster*, *Lasius*, *Messor*, *Monomorium*, *Pheidole*, *Plagiopileps*, *Temnothorax* and *Tetramorium*. The number of genera per site varied between 3 and 7 including *Tapinoma* (mean = 4.2).

## Differences in activity between *T. magnum* and *T. darioi

The GAMMs of *T. magnum* and *T. darioi* representing the average abundance score as a function of temperature explain 46.4% and 48.3% of the deviance respectively (Figure S1). For *T. magnum*, temperature and site had a significant effect on the average abundance score (*p* < 0.001 and *p* = 0.01). For *T. darioi*, temperature and site had
<figure><img src="image_7.png" /><figcaption>**FIGURE 3** Generalised additive mixed model predictions of *Tapinoma magnum* workers (green dotted line, *n* = 76 measurement sessions) and *Tapinoma darioi* workers (orange solid line, *n* = 71 measurement sessions) average abundance scores, with 95% confidence intervals, according to (a) temperature and (b) day of the year (black vertical bars above the x-axis represent sampling days and coloured disk the start of a season). The abundance score of ants is classified into five categories corresponding to the average number of individuals on the 18 baits (0 = 0 worker; 1 = 1 worker; 2 = 2-5 workers; 3 = 6-20 workers; 4 = 21-100 workers; 5 = more than 100 workers). The grey area represents the time period when the activities of *T. magnum* and *T. darioi* differed significantly.</figcaption></figure>

(GAMMs) (Wood, 2006). This semi-parametric model is a flexible regression method for estimating linear and non-linear trends. GAMMs are often used to predict smooth functional relationships between the response variable and the explanatory variables, with the advantage that there are no predefined shape for describing functional relationships (Pedersen et al., 2019; Wood, 2017). Additionally, the foraging activity measurements followed an unevenly spaced time series because the measurement sessions were irregularly spaced in time, which made it difficult to apply conventional time series methods (Simpson, 2018). According to the methods of Andersen (1992), an abundance score (0 = 0 worker; 1 = 1 worker; 2 = 2-5 workers; 3 = 6-20 workers; 4 = 21-100 workers; 5 = more than 100 workers) was calculated for each bait occupied by the ant species considered. Then the average abundance score on the 18 baits was calculated and used as the foraging activity measurement because using the number of individuals recorded yielded overdispersed residuals in the models.

First, independent GAMMs were parameterised for *T. magnum* and for *T. darioi*, respectively, including as response variable the average abundance score and as explanatory variable the mean temperature with the smooth thin plate regression splines term and the site as random effect. Thin plate regression splines can be used to model complex non-linear relationships between variables, smoothing the data efficiently and flexibly without having to specify the location of the nodes (Wood, 2003). Two new GAMMs (one for each species) were parameterised with the same variables, but replacing the mean temperature by the day of the year as explanatory variable with the smooth cyclic cubic spline term. This smoothing term forces the ends of the spline to be equal to each other, allowing continuity between the first and last day of the year (Wood, 2017). In a second phase, the same GAMMs were carried out, considering the activities of both

invasive *Tapinoma* species together to infer their general activity pattern and compare it with that of native species.

This approach provides a reliable estimate by using 95% confidence limits around the modelled responses which we used to determine the differences in activity between ants in different categories (i.e., *T. magnum* vs. *T. darioi* and *Tapinoma* vs. native ants). For all the models, the distribution of the specified errors was quasi-Poisson. Model parameters were chosen by restricted maximum likelihood (REML) selection. All the GAMMs were produced using the ‘gam’ function in the ‘mgcv’ package (Wood, 2011) using R 4.2.1 software (R Core Team, 2022) and R studio (RStudio Core Team, 2020).

## RESULTS

### Species detected on baits

Overall, nine genera other than *Tapinoma* were recorded on the baits: *Camponotus*, *Crematogaster*, *Lasius*, *Messor*, *Monomorium*, *Pheidole*, *Plagiopileps*, *Temnothorax* and *Tetramorium*. The number of genera per site varied between 3 and 7 including *Tapinoma* (mean = 4.2).

### Differences in activity between *T. magnum* and *T. darioi

The GAMMs of *T. magnum* and *T. darioi* representing the average abundance score as a function of temperature explain 46.4% and 48.3% of the deviance respectively (Figure S1). For *T. magnum*, temperature and site had a significant effect on the average abundance score (*p* < 0.001 and *p* = 0.01). For *T. darioi*, temperature and site had
a significant effect on the average abundance score (p < 0.001 and p = 0.02). Along the observed temperature gradient, the average abundance score did not differ significantly between *T. magnum* and *T. darioi*. The maximum levels of activity for *T. magnum* and *T. darioi* were 3.2 and 1.7 of average abundance score, respectively, although these maximum levels of activity were not significantly different (Figure 3a). The activity of both species was highest between 18 and 20°C.

The GAMMs of *T. magnum* and *T. darioi* representing the average abundance score according to the day of the year explain 55.5% and 74.4% of the deviance respectively. For *T. magnum*, the day of the year had a significant effect on the average abundance score (p < 0.001), but the site did not (p = 0.07). For *T. darioi*, the day of year and site both had a significant effect on the average abundance score (p < 0.001 and p = 0.001, respectively). The average abundance score differed significantly between *T. magnum* and *T. darioi* over the period from day 340 to day 365 and from day 1 to day 40 of the year (early December to mid-February, i.e., autumn-winter). Over this period, *T. darioi* had an average abundance score close to 0. In contrast, *T. magnum* was active throughout the year with a minimum average abundance score at 0.5. The average abundance score of *T. magnum* and *T. darioi* began to increase from day 340 (early December, i.e., autumn) and day 30 (early February, i.e., winter) of the year, respectively. Both species reached their peak activity on day 100 (mid-April, i.e., spring), and then the average abundance score decreased throughout the year, however, the peak activity of *T. magnum* (3.6) was not significantly higher than that of *T. darioi* (1.6) (Figure 3b).

## Differences in activity between *Tapinoma* and the native ant species

The GAMMs for *Tapinoma* and native species, representing the average abundance score as a function of temperature, explain 52.3% and 90.2% of the deviance, respectively (Figure S1). For *Tapinoma* and native species, temperature and site had a significant effect on the average abundance score (p < 0.001 in both cases). Along the temperature gradient observed, the average abundance score by *Tapinoma* was significantly higher than that of native species for the range of lowest temperatures (6–19°C). *Tapinoma* activity increased strongly from 3 to 20°C where it reached a maximum (2.3 of average abundance score) and then decreased, while native species activity increased slowly from 7 to 27°C where it reached a maximum (1 of average abundance score) and then decreased. The two peaks of activity are significantly different (Figure 4a).

The GAMMs for *Tapinoma* and native species representing the average abundance score according to the day of the year explain 66.9% and 88.8% of the deviance respectively. For *Tapinoma* and native species, day of year and site had a significant effect on the average abundance score (p < 0.001 in both cases). The average abundance score by *Tapinoma* was significantly higher than that of native species over the period from day 30 to day 120 of the year

(early February to end April, i.e., winter-spring). The average abundance score by *Tapinoma* increased from day 10 to day 100 (mid-January to mid-April, i.e., winter-spring), when it peaked (2.3 of average abundance score), and then decreased. The average abundance score by native ants increased from day 70 to day 250 (mid-March to early September, i.e., spring-summer), when it reached a maximum (1.1), and then decreased. The two activity peaks are significantly different (Figure 4b).

## DISCUSSION

In this study, we show how temperature and season regulate the activities of *T. darioi* and *T. magnum*, as well as native species in an invaded Mediterranean environment. We observed a similar range of temperature of activity between *T. darioi* and *T. magnum*. However, *T. magnum* was more active than *T. darioi* in winter, which may indicate the influence of seasonal factors other than temperature. We were able to identify differences in activity between *Tapinoma* and native species depending on the temperature or the season following in situ activity measurements throughout the year.

## Establishment success of *T. magnum* and *T. darioi*

Temperature is a major determinant of the distribution and ability of ants to establish new populations throughout the world (Jung et al., 2021; Roeder et al., 2018; Warren II & Chick, 2013). It has also been established that the temperature-dependent activity niche plays a crucial role in ant distribution, even superseding the importance of critical thermal limits in some cases (Guo et al., 2020). Our results highlighted *Tapinoma* activity under a wide range of temperatures, including relatively cold temperatures, potentially explaining the establishment of these species in regions that are colder than their native range (Dekoninck et al., 2015; Noordijk, 2016; Seifert et al., 2017). Furthermore, Noordijk (2016) also observed sunbathing activity of these *Tapinoma* in winter in the Netherlands, which confirm that these ants are active at cold temperatures even in invaded northern European locations. With ongoing climate change and its positive impact on the spread of invasive ants (Bertelsmeier et al., 2015, 2016), these two species represent a major threat on a very large scale.

In this study, we did not detect any difference in activity between the two species according to temperature. According to our results, the differences in distribution between the two species in Western Europe (Gouraud & Kaufmann, 2022; Lenoir et al., 2023; Seifert et al., 2017) may therefore not be explained by foraging temperature preferences. Nevertheless, we observed differences in activity between the two species in winter that were not explained by temperature. Other climatic parameters involved in seasonality, such as humidity or photoperiod, are known to influence ant activity (KimFung & ChowYang, 2009; Lei et al., 2019) and may play a role in the establishment of these species. However, due to the limited
a significant effect on the average abundance score ($p$ < 0.001 and $p$ = 0.02). Along the observed temperature gradient, the average abundance score did not differ significantly between *T. magnum* and *T. darioi*. The maximum levels of activity for *T. magnum* and *T. darioi* were 3.2 and 1.7 of average abundance score, respectively, although these maximum levels of activity were not significantly different (Figure 3a). The activity of both species was highest between 18 and 20°C.

The GAMMs of *T. magnum* and *T. darioi* representing the average abundance score according to the day of the year explain 55.5% and 74.4% of the deviance respectively. For *T. magnum*, the day of the year had a significant effect on the average abundance score ($p$ < 0.001), but the site did not ($p$ = 0.07). For *T. darioi*, the day of year and site both had a significant effect on the average abundance score ($p$ < 0.001 and $p$ = 0.001, respectively). The average abundance score differed significantly between *T. magnum* and *T. darioi* over the period from day 340 to day 365 and from day 1 to day 40 of the year (early December to mid-February, i.e., autumn-winter). Over this period, *T. darioi* had an average abundance score close to 0. In contrast, *T. magnum* was active throughout the year with a minimum average abundance score at 0.5. The average abundance score of *T. magnum* and *T. darioi* began to increase from day 340 (early December, i.e., autumn) and day 30 (early February, i.e., winter) of the year, respectively. Both species reached their peak activity on day 100 (mid-April, i.e., spring), and then the average abundance score decreased throughout the year, however, the peak activity of *T. magnum* (3.6) was not significantly higher than that of *T. darioi* (1.6) (Figure 3b).

## Differences in activity between *Tapinoma* and the native ant species

The GAMMs for *Tapinoma* and native species, representing the average abundance score as a function of temperature, explain 52.3% and 90.2% of the deviance, respectively (Figure S1). For *Tapinoma* and native species, temperature and site had a significant effect on the average abundance score ($p$ < 0.001 in both cases). Along the temperature gradient observed, the average abundance score by *Tapinoma* was significantly higher than that of native species for the range of lowest temperatures (6-19°C). *Tapinoma* activity increased strongly from 3 to 20°C where it reached a maximum (2.3 of average abundance score) and then decreased, while native species activity increased slowly from 7 to 27°C where it reached a maximum (1 of average abundance score) and then decreased. The two peaks of activity are significantly different (Figure 4a).

The GAMMs for *Tapinoma* and native species representing the average abundance score according to the day of the year explain 66.9% and 88.8% of the deviance respectively. For *Tapinoma* and native species, day of year and site had a significant effect on the average abundance score ($p$ < 0.001 in both cases). The average abundance score by *Tapinoma* was significantly higher than that of native species over the period from day 30 to day 120 of the year

(early February to end April, i.e., winter-spring). The average abundance score by *Tapinoma* increased from day 10 to day 100 (mid-January to mid-April, i.e., winter-spring), when it peaked (2.3 of average abundance score), and then decreased. The average abundance score by native ants increased from day 70 to day 250 (mid-March to early September, i.e., spring-summer), when it reached a maximum (1.1), and then decreased. The two activity peaks are significantly different (Figure 4b).

## DISCUSSION

In this study, we show how temperature and season regulate the activities of *T. darioi* and *T. magnum*, as well as native species in an invaded Mediterranean environment. We observed a similar range of temperature of activity between *T. darioi* and *T. magnum*. However, *T. magnum* was more active than *T. darioi* in winter, which may indicate the influence of seasonal factors other than temperature. We were able to identify differences in activity between *Tapinoma* and native species depending on the temperature or the season following in situ activity measurements throughout the year.

## Establishment success of *T. magnum* and *T. darioi*

Temperature is a major determinant of the distribution and ability of ants to establish new populations throughout the world (Jung et al., 2021; Roeder et al., 2018; Warren II & Chick, 2013). It has also been established that the temperature-dependent activity niche plays a crucial role in ant distribution, even superseding the importance of critical thermal limits in some cases (Guo et al., 2020). Our results highlighted *Tapinoma* activity under a wide range of temperatures, including relatively cold temperatures, potentially explaining the establishment of these species in regions that are colder than their native range (Dekoninck et al., 2015; Noordijk, 2016; Seifert et al., 2017). Furthermore, Noordijk (2016) also observed sunbathing activity of these *Tapinoma* in winter in the Netherlands, which confirm that these ants are active at cold temperatures even in invaded northern European locations. With ongoing climate change and its positive impact on the spread of invasive ants (Bertelsmeier et al., 2015, 2016), these two species represent a major threat on a very large scale.

In this study, we did not detect any difference in activity between the two species according to temperature. According to our results, the differences in distribution between the two species in Western Europe (Gouraud & Kaufmann, 2022; Lenoir et al., 2023; Seifert et al., 2017) may therefore not be explained by foraging temperature preferences. Nevertheless, we observed differences in activity between the two species in winter that were not explained by temperature. Other climatic parameters involved in seasonality, such as humidity or photoperiod, are known to influence ant activity (KimFung & ChowYang, 2009; Lei et al., 2019) and may play a role in the establishment of these species. However, due to the limited
<figure><img src="image_8.png" /></figure>

<figure><img src="image_9.png" /><figcaption>**FIGURE 4** Generalised additive mixed model predictions of *Tapinoma* workers (red dotted line) and native ant workers (blue solid line) (n = 147 measurement sessions) average abundance scores, with 95% confidence intervals, according to (a) temperature and (b) day of the year (black vertical bars above the x-axis represent sampling days and coloured disk the start of a season). The abundance score of ants is classified into five categories corresponding to the average number of individuals on the 18 baits (0 = 0 worker; 1 = 1 worker; 2 = 2-5 workers; 3 = 6-20 workers; 4 = 21-100 workers; 5 = more than 100 workers). The grey area represents the temperature range and the time period when the activities of *Tapinoma* and native species differed significantly.</figcaption></figure>

number of colonies of each *Tapinoma* species observed, we must remain cautious about these results.

Our results support the hypothesis that climatic parameters such as temperature have little influence on the difference in establishment between these species in continental Europe. Biotic factors, such as competition, could better explain the difference in establishment between *T. magnum* and *T. darioi* depending on locality (de Soares, 2013; Ribas & Schoereder, 2002). These two species seem to share similar temperature niches and type of environment (anthropised environment) (Centanni et al., 2022; Gouraud & Kaufmann, 2022), so they may be strong competitors. Thus, the difference in establishment between the two invasive *Tapinoma* species in non-native areas could have more to do with differential introduction patterns (Destour et al., in prep). One hypothesis would be that the species that is established into a given environment would be the species that has been introduced first, as it would have no competitors in the invaded environment and would then exclude the other by competitive exclusion (priority effect) (Cleland et al., 2015; Fukami, 2015; Weidlich et al., 2021). In addition, Palmer et al. (2002, 2003) demonstrated that in ants, priority effect can allow an ant species to avoid the exclusion by another ant species whatever the competitive hierarchy between them. On the other hand, *T. magnum* is more active in winter than *T. darioi*. This species therefore appears to be better able to compete with non-Mediterranean species than *T. darioi*, which could facilitate its establishment in northern Europe.

## Niche partitioning between *Tapinoma* and native ants

We observed that the peak activity of *Tapinoma* occurred at a lower temperature and earlier in the season from that of native species.

This difference, which allows different activity niches to be occupied, may explain the coexistence of *Tapinoma* and native species (Gippet et al., 2022). For the origin of this difference in activity periods, we know that workers of the *T. nigerrimum* complex are generally active relatively early in the year compared with other Mediterranean species (Carpintero et al., 2007; Cros et al., 1997). The swarming period of these *Tapinoma* species starts at the end of spring, so their colonies probably need more resources just before this period than species with a later swarming period. The other Mediterranean species observed in this study have a swarming period between late spring and early autumn (Blatrix et al., 2013). The period of peak activity of *Tapinoma* species, which is initially out of sync with other species, may be a factor that facilitates their establishment in invaded areas. In fact, such a phenology could enable them to avoid competition with dominant native species able to limit the establishment of invasive species (Cordonnier et al., 2020; Krushelnycky & Gillespie, 2010; Wetterer et al., 2006). *Tapinoma magnum* has the ability to shift its thermal niche (Bujan et al., 2021), potentially accentuating this difference in activity periods (Broennimann et al., 2007). We cannot tell from our results whether the timing of the peak activity we observed for *T. magnum* was intrinsic to the species, or induced or accentuated by a niche shift following its introduction. The difference in activity observed between *Tapinoma* and native ants may also be due to a shift in the activity period of native species caused by the presence of *Tapinoma*. Furthermore, the difference in activity periods, due to temporal heterogeneity, may not be the only difference that allows coexistence of native ants and *Tapinoma* species. Gippet et al. (2022) showed that heterogeneity in microclimatic conditions could also play an important role in coexistence in areas invaded by *T. magnum*.
<figure><img src="image_8.png" /></figure>

<figure><img src="image_9.png" /><figcaption>**FIGURE 4** Generalised additive mixed model predictions of *Tapinoma* workers (red dotted line) and native ant workers (blue solid line) (n = 147 measurement sessions) average abundance scores, with 95% confidence intervals, according to (a) temperature and (b) day of the year (black vertical bars above the x-axis represent sampling days and coloured disk the start of a season). The abundance score of ants is classified into five categories corresponding to the average number of individuals on the 18 baits (0 = 0 worker; 1 = 1 worker; 2 = 2-5 workers; 3 = 6-20 workers; 4 = 21-100 workers; 5 = more than 100 workers). The grey area represents the temperature range and the time period when the activities of *Tapinoma* and native species differed significantly.</figcaption></figure>

number of colonies of each *Tapinoma* species observed, we must remain cautious about these results.

Our results support the hypothesis that climatic parameters such as temperature have little influence on the difference in establishment between these species in continental Europe. Biotic factors, such as competition, could better explain the difference in establishment between *T. magnum* and *T. darioi* depending on locality (de Soares, 2013; Ribas & Schoereder, 2002). These two species seem to share similar temperature niches and type of environment (anthropised environment) (Centanni et al., 2022; Gouraud & Kaufmann, 2022), so they may be strong competitors. Thus, the difference in establishment between the two invasive *Tapinoma* species in non-native areas could have more to do with differential introduction patterns (Destour et al., in prep). One hypothesis would be that the species that is established into a given environment would be the species that has been introduced first, as it would have no competitors in the invaded environment and would then exclude the other by competitive exclusion (priority effect) (Cleland et al., 2015; Fukami, 2015; Weidlich et al., 2021). In addition, Palmer et al. (2002, 2003) demonstrated that in ants, priority effect can allow an ant species to avoid the exclusion by another ant species whatever the competitive hierarchy between them. On the other hand, *T. magnum* is more active in winter than *T. darioi*. This species therefore appears to be better able to compete with non-Mediterranean species than *T. darioi*, which could facilitate its establishment in northern Europe.

## Niche partitioning between *Tapinoma* and native ants

We observed that the peak activity of *Tapinoma* occurred at a lower temperature and earlier in the season from that of native species.

This difference, which allows different activity niches to be occupied, may explain the coexistence of *Tapinoma* and native species (Gippet et al., 2022). For the origin of this difference in activity periods, we know that workers of the *T. nigerrimum* complex are generally active relatively early in the year compared with other Mediterranean species (Carpintero et al., 2007; Cros et al., 1997). The swarming period of these *Tapinoma* species starts at the end of spring, so their colonies probably need more resources just before this period than species with a later swarming period. The other Mediterranean species observed in this study have a swarming period between late spring and early autumn (Blatrix et al., 2013). The period of peak activity of *Tapinoma* species, which is initially out of sync with other species, may be a factor that facilitates their establishment in invaded areas. In fact, such a phenology could enable them to avoid competition with dominant native species able to limit the establishment of invasive species (Cordonnier et al., 2020; Krushelnycky & Gillespie, 2010; Wetterer et al., 2006). *Tapinoma magnum* has the ability to shift its thermal niche (Bujan et al., 2021), potentially accentuating this difference in activity periods (Broennimann et al., 2007). We cannot tell from our results whether the timing of the peak activity we observed for *T. magnum* was intrinsic to the species, or induced or accentuated by a niche shift following its introduction. The difference in activity observed between *Tapinoma* and native ants may also be due to a shift in the activity period of native species caused by the presence of *Tapinoma*. Furthermore, the difference in activity periods, due to temporal heterogeneity, may not be the only difference that allows coexistence of native ants and *Tapinoma* species. Gippet et al. (2022) showed that heterogeneity in microclimatic conditions could also play an important role in coexistence in areas invaded by *T. magnum*.
## Management implications

Tapinoma is spreading rapidly, particularly through the transport of ornamental plants, and more and more populations are being established in Europe (Gouraud & Kaufmann, 2022; Lenoir et al., 2023; Seifert et al., 2017). Given its negative impact on local ants and the potential damage caused by the formation of supercolonies (Gippet et al., 2022; Helanterä, 2022), it is becoming essential to eradicate or at least control populations established in non-native areas. However, the eradication of invasive ants often requires the use of toxic baits, representing a danger for the native fauna of the invaded locality (Hoffmann et al., 2016). To minimise the impact of these toxic baits, we should consider applying them during periods when native ants are less active, or even inactive (Hoffmann et al., 2011). Planning an eradication for late winter could be beneficial for the effective elimination of Tapinoma, as their activity peaks at this time, sparing other native ants that are less active during this period in Mediterranean areas.

However, we should be aware that the eradication success rate for ants is relatively low (66%), that is 20% lower than for invasive mammals on islands (Hoffmann et al., 2016). Therefore, it is advisable to focus on managing invasion routes rather than investing in eradication attempts (Hulme, 2009; Pergl et al., 2017).

## AUTHOR CONTRIBUTIONS

**Giovanny Destour:** Data curation; formal analysis; writing – original draft; writing – review and editing.
**Rumsais Blatrix:** Supervision; conceptualization; methodology; writing – review and editing; funding acquisition; formal analysis; data curation.
**Aurélien Caries:** Writing – review and editing; investigation.
**Juliette Genevet:** Investigation; writing – review and editing.
**Bernard Kaufmann:** Conceptualization; writing – review and editing; data curation; methodology.
**Baptiste Lecoq:** Investigation; writing – review and editing.
**Annick Lucas:** Investigation; writing – review and editing.
**Killian Verlingue:** Investigation; writing – review and editing.
**Alan Vergnes:** Conceptualization; writing – review and editing; formal analysis; data curation; supervision; funding acquisition; methodology.
**Marion Javal:** Conceptualization; methodology; funding acquisition; writing – review and editing; formal analysis; supervision; project administration.

## ACKNOWLEDGEMENTS

The authors thank Alan Andersen and François Brassard for their comments on an earlier version of the manuscript and language edition. The authors thank two anonymous reviewers for their very useful comments, which helped to improve the manuscript.

## FUNDING INFORMATION

This study was funded by Montpellier Méditerranée Metropolis and the City of Montpellier as part of their agreement with CEFE (2021–2026).

## CONFLICT OF INTEREST STATEMENT

The authors declare no conflict of interest.

## DATA AVAILABILITY STATEMENT

The data that support the findings of this study are available in the supplementary material of this article.

## ETHICS STATEMENT

The authors have respected ethical standards in the preparation of this article.

## ORCID

Giovanny Destour https://orcid.org/0009-0002-5725-5816

Rumsais Blatrix https://orcid.org/0000-0003-1662-7791

Aurélien Caries https://orcid.org/0009-0002-9507-7475

Bernard Kaufmann https://orcid.org/0000-0001-9097-3452

Baptiste Lecoq https://orcid.org/0009-0005-2834-9436

Alan Vergnes https://orcid.org/0000-0002-6128-0793

Marion Javal https://orcid.org/0000-0001-7878-2936

## REFERENCES

Albrecht, M. & Gotelli, N.J. (2001) Spatial and temporal niche partitioning in grassland ants. *Oecologia*, 126, 134–141.

Andersen, A.N. (1992) Regulation of “momentary” diversity by dominant species in exceptionally rich ant communities of the Australian seasonal tropics. *The American Naturalist*, 140, 401–420.

Angilletta, M.J. (2009) *Thermal adaptation: a theoretical and empirical synthesis*. New York: Oxford University Press.

Bates, O.K. & Bertelsmeier, C. (2021) Climatic niche shifts in introduced species. *Current Biology*, 31, R1252–R1266.

Bertelsmeier, C., Blight, O. & Courchamp, F. (2016) Invasions of ants (Hymenoptera: Formicidae) in light of global climate change. *Myrmecological News*, 22, 25–42.

Bertelsmeier, C., Luque, G.M., Hoffmann, B.D. & Courchamp, F. (2015) Worldwide ant invasions under climate change. *Biodiversity and Conservation*, 24, 117–128.

Bertelsmeier, C., Ollier, S., Liebhold, A.M., Brockerhoff, E.G., Ward, D. & Keller, L. (2018) Recurrent bridgehead effects accelerate global alien ant spread. *Proceedings of the National Academy of Sciences*, 115, 5486–5491.

Berthon, K. (2015) How do native species respond to invaders? Mechanistic and trait-based perspectives. *Biological Invasions*, 17, 2199–2211.

Blatrix, R., Galkowski, C., Lebas, C. & Wegnez, P. (2013) *Fourmis de France, de Belgique et du Luxembourg*. Luçon, Delachaux et Niestlé, 288 pp.

Blight, O., Provost, E., Renucci, M., Tirard, A. & Orgeas, J. (2010) A native ant armed to limit the spread of the Argentine ant. *Biological Invasions*, 12, 3785–3793.

Brinkman, M.A., Gardner, W.A., Ipser, R.M. & Diffie, S.K. (2001) Ground-dwelling ant species attracted to four food baits in Georgia. *Journal of Entomological Science*, 36, 461–463.

Broennimann, O., Treier, U.A., Müller-Schärer, H., Thuiller, W., Peterson, A.T. & Guisan, A. (2007) Evidence of climatic niche shift during biological invasion. *Ecology Letters*, 10, 701–709.

Bujan, J., Charavel, E., Bates, O.K., Gippet, J.M.W., Darras, H., Lebas, C. et al. (2021) Increased acclimation ability accompanies a thermal niche shift of a recent invasion. *Journal of Animal Ecology*, 90, 483–491.

Carpintero, S., Retana, J., Cerdá, X., Reyes-López, J. & Arias de Reyna, L. (2007) Exploitative strategies of the invasive Argentine ant (*Linepithema humile*) and native ant species in a southern Spanish pine forest. *Environmental Entomology*, 36, 1100–1111.

Centanni, J., Kaufmann, B., Blatrix, R., Blight, O., Dumet, A., Jay-Robert, P. et al. (2022) High resolution mapping in Southern France reveals that distributions of supercolonial and monodomous species in the
## Management implications

Tapinoma is spreading rapidly, particularly through the transport of ornamental plants, and more and more populations are being established in Europe (Gouraud & Kaufmann, 2022; Lenoir et al., 2023; Seifert et al., 2017). Given its negative impact on local ants and the potential damage caused by the formation of supercolonies (Gippet et al., 2022; Helanterä, 2022), it is becoming essential to eradicate or at least control populations established in non-native areas. However, the eradication of invasive ants often requires the use of toxic baits, representing a danger for the native fauna of the invaded locality (Hoffmann et al., 2016). To minimise the impact of these toxic baits, we should consider applying them during periods when native ants are less active, or even inactive (Hoffmann et al., 2011). Planning an eradication for late winter could be beneficial for the effective elimination of Tapinoma, as their activity peaks at this time, sparing other native ants that are less active during this period in Mediterranean areas.

However, we should be aware that the eradication success rate for ants is relatively low (66%), that is 20% lower than for invasive mammals on islands (Hoffmann et al., 2016). Therefore, it is advisable to focus on managing invasion routes rather than investing in eradication attempts (Hulme, 2009; Pergl et al., 2017).

## AUTHOR CONTRIBUTIONS

**Giovanny Destour:** Data curation; formal analysis; writing – original draft; writing – review and editing.
**Rumsais Blatrix:** Supervision; conceptualization; methodology; writing – review and editing; funding acquisition; formal analysis; data curation.
**Aurélien Caries:** Writing – review and editing; investigation.
**Juliette Genevet:** Investigation; writing – review and editing.
**Bernard Kaufmann:** Conceptualization; writing – review and editing; data curation; methodology.
**Baptiste Lecoq:** Investigation; writing – review and editing.
**Annick Lucas:** Investigation; writing – review and editing.
**Killian Verlingue:** Investigation; writing – review and editing.
**Alan Vergnes:** Conceptualization; writing – review and editing; formal analysis; data curation; supervision; funding acquisition; methodology.
**Marion Javal:** Conceptualization; methodology; funding acquisition; writing – review and editing; formal analysis; supervision; project administration.

## ACKNOWLEDGEMENTS

The authors thank Alan Andersen and François Brassard for their comments on an earlier version of the manuscript and language edition. The authors thank two anonymous reviewers for their very useful comments, which helped to improve the manuscript.

## FUNDING INFORMATION

This study was funded by Montpellier Méditerranée Metropolis and the City of Montpellier as part of their agreement with CEFE (2021–2026).

## CONFLICT OF INTEREST STATEMENT

The authors declare no conflict of interest.

## DATA AVAILABILITY STATEMENT

The data that support the findings of this study are available in the supplementary material of this article.

## ETHICS STATEMENT

The authors have respected ethical standards in the preparation of this article.

## ORCID

Giovanny Destour https://orcid.org/0009-0002-5725-5816

Rumsais Blatrix https://orcid.org/0000-0003-1662-7791

Aurélien Caries https://orcid.org/0009-0002-9507-7475

Bernard Kaufmann https://orcid.org/0000-0001-9097-3452

Baptiste Lecoq https://orcid.org/0009-0005-2834-9436

Alan Vergnes https://orcid.org/0000-0002-6128-0793

Marion Javal https://orcid.org/0000-0001-7878-2936

## REFERENCES

Albrecht, M. & Gotelli, N.J. (2001) Spatial and temporal niche partitioning in grassland ants. *Oecologia*, 126, 134–141.

Andersen, A.N. (1992) Regulation of “momentary” diversity by dominant species in exceptionally rich ant communities of the Australian seasonal tropics. *The American Naturalist*, 140, 401–420.

Angilletta, M.J. (2009) *Thermal adaptation: a theoretical and empirical synthesis*. New York: Oxford University Press.

Bates, O.K. & Bertelsmeier, C. (2021) Climatic niche shifts in introduced species. *Current Biology*, 31, R1252–R1266.

Bertelsmeier, C., Blight, O. & Courchamp, F. (2016) Invasions of ants (Hymenoptera: Formicidae) in light of global climate change. *Myrmecological News*, 22, 25–42.

Bertelsmeier, C., Luque, G.M., Hoffmann, B.D. & Courchamp, F. (2015) Worldwide ant invasions under climate change. *Biodiversity and Conservation*, 24, 117–128.

Bertelsmeier, C., Ollier, S., Liebhold, A.M., Brockerhoff, E.G., Ward, D. & Keller, L. (2018) Recurrent bridgehead effects accelerate global alien ant spread. *Proceedings of the National Academy of Sciences*, 115, 5486–5491.

Berthon, K. (2015) How do native species respond to invaders? Mechanistic and trait-based perspectives. *Biological Invasions*, 17, 2199–2211.

Blatrix, R., Galkowski, C., Lebas, C. & Wegnez, P. (2013) *Fourmis de France, de Belgique et du Luxembourg*. Luçon, Delachaux et Niestlé, 288 pp.

Blight, O., Provost, E., Renucci, M., Tirard, A. & Orgeas, J. (2010) A native ant armed to limit the spread of the Argentine ant. *Biological Invasions*, 12, 3785–3793.

Brinkman, M.A., Gardner, W.A., Ipser, R.M. & Diffie, S.K. (2001) Ground-dwelling ant species attracted to four food baits in Georgia. *Journal of Entomological Science*, 36, 461–463.

Broennimann, O., Treier, U.A., Müller-Schärer, H., Thuiller, W., Peterson, A.T. & Guisan, A. (2007) Evidence of climatic niche shift during biological invasion. *Ecology Letters*, 10, 701–709.

Bujan, J., Charavel, E., Bates, O.K., Gippet, J.M.W., Darras, H., Lebas, C. et al. (2021) Increased acclimation ability accompanies a thermal niche shift of a recent invasion. *Journal of Animal Ecology*, 90, 483–491.

Carpintero, S., Retana, J., Cerdá, X., Reyes-López, J. & Arias de Reyna, L. (2007) Exploitative strategies of the invasive Argentine ant (*Linepithema humile*) and native ant species in a southern Spanish pine forest. *Environmental Entomology*, 36, 1100–1111.

Centanni, J., Kaufmann, B., Blatrix, R., Blight, O., Dumet, A., Jay-Robert, P. et al. (2022) High resolution mapping in Southern France reveals that distributions of supercolonial and monodomous species in the
*Tapinoma nigerrimum* complex (Hymenoptera: Formicidae) are related to sensitivity to urbanization. *Myrmecological News*, 32, 41–50.

Cleland, E.E., Esch, E. & McKinney, J. (2015) Priority effects vary with species identity and origin in an experiment varying the timing of seed arrival. *Oikos*, 124, 33–40.

Cordonnier, M., Blight, O., Angulo, E. & Courchamp, F. (2020) The native ant *Lasius niger* can limit the access to resources of the invasive Argentine ant. *Animals*, 10, 2451.

Cros, S., Cerdá, X. & Retana, J. (1997) Spatial and temporal variations in the activity patterns of Mediterranean ant communities. *Écoscience*, 4, 269–278.

Davidson, D.W. (1998) Resource discovery versus resource domination in ants: a functional mechanism for breaking the trade-off. *Ecological Entomology*, 23, 484–490.

de Soares, S.A. (2013) The role of competition in structuring ant communities: a review. *Oecologia Australis*, 17, 271–281.

Dekoninck, W., Parmentier, T. & Seifert, B. (2015) First records of a super-colonial species of the *Tapinoma nigerrimum* complex in Belgium (Hymenoptera: Formicidae). *Bulletin van de Koninklijke Belgische Vereniging voor Entomologie*, 151, 206–209.

Diagne, C., Leroy, B., Vaissière, A.-C., Gozlan, R.E., Roiz, D., Jarić, I. et al. (2021) High and rising economic costs of biological invasions worldwide. *Nature*, 592, 571–576.

Didham, R.K., Tylianakis, J.M., Hutchison, M.A., Ewers, R.M. & Gemmell, N.J. (2005) Are invasive species the drivers of ecological change? *Trends in Ecology & Evolution*, 20, 470–474.

Dijkstra, J.A., Westerman, E.L. & Harris, L.G. (2017) Elevated seasonal temperatures eliminate thermal barriers of reproduction of a dominant invasive species: a community state change for northern communities? *Diversity and Distributions*, 23, 1182–1192.

Fukami, T. (2015) Historical contingency in community assembly: integrating niches, species pools, and priority effects. *Annual Review of Ecology, Evolution, and Systematics*, 46, 1–23.

Gan, J.Y.S., Pathy, T.S., Yeo, D.Y.C., Lee, J.M. & Yek, S.H. (2022) The population dynamics of the yellow crazy ant *Anoplolepis gracilipes* (Hymenoptera: Formicidae) on a tropical Island in Malaysia. *Ecological Entomology*, 47, 339–346.

Gippet, J.M.W., George, L. & Bertelsmeier, C. (2022) Local coexistence of native and invasive ant species is associated with micro-spatial shifts in foraging activity. *Biological Invasions*, 24, 761–773.

Gouraud, C. & Kaufmann, B. (2022) Nouvelles observations des fourmis invasives du complexe des *Tapinoma* gr. *nigerrimum* dans le Massif armoricain (Hymenoptera: Formicidae). *Invertébrés Armoricains*, 23, 23–38.

Guo, F., Guénard, B., Economo, E.P., Deutsch, C.A. & Bonebrake, T.C. (2020) Activity niches outperform thermal physiological limits in predicting global ant distributions. *Journal of Biogeography*, 47, 829–842.

Helanterä, H. (2022) Supercolonies of ants (Hymenoptera: Formicidae): ecological patterns, behavioural processes and their implications for social evolution. *Myrmecological News*, 32, 1–22.

Hoffmann, B., Davis, P., Gott, K., Jennings, C., Joe, S., Krushelnycky, P. et al. (2011) Improving ant eradications: details of more successes, a global synthesis and recommendations. *Aliens*, 31, 16–23.

Hoffmann, B.D., Luque, G.M., Bellard, C., Holmes, N.D. & Donlan, C.J. (2016) Improving invasive ant eradication as a conservation tool: a review. *Biological Conservation*, 198, 37–49.

Holway, D.A., Lach, L., Suarez, A.V., Tsutsui, N.D. & Case, T.J. (2002) The causes and consequences of ant invasions. *Annual Review of Ecology and Systematics*, 33, 181–233.

Hulme, P.E. (2009) Trade, transport and trouble: managing invasive species pathways in an era of globalization. *Journal of Applied Ecology*, 46, 10–18.

INSEE. (2022). https://www.insee.fr/fr/statistiques/1405599?geo=DEP-34+COM-34172. Accessed on 13 May 2024.

Jung, J.-M., Lee, H.-S., Lee, J.-H., Jung, S. & Lee, W.-H. (2021) Development of a predictive model for soil temperature and its application to species distribution modeling of ant species in South Korea. *Ecological Informatics*, 61, 101220.

KimFung, C. & ChowYang, L. (2009) Influences of temperature, relative humidity and light intensity on the foraging activity of field populations of the longlegged ant, *Anoplolepis gracilipes* (Hymenoptera: Formicidae). *Sociobiology*, 54, 531–540.

Krushelnycky, P.D. & Gillespie, R.G. (2010) Correlates of vulnerability among arthropod species threatened by invasive ants. *Biodiversity and Conservation*, 19, 1971–1988.

Lei, Y., Zhou, Y., Lü, L. & He, Y. (2019) Rhythms in foraging behavior and expression patterns of the foraging gene in *Solenopsis invicta* (Hymenoptera: Formicidae) in relation to photoperiod. *Journal of Economic Entomology*, 112, 2923–2930.

Lenoir, A., Mercier, J.-L., Perdereau, E., Berville, L. & Galkowski, C. (2023) Sur l'expansion des fourmis envahissantes du genre *Tapinoma* en France (Hymenoptera: Formicidae). *Osmia*, 11, 1–10.

Lord, J.P. (2017) Impact of seawater temperature on growth and recruitment of invasive fouling species at the global scale. *Marine Ecology*, 38, e12404.

Lowe, S., Browne, M., Boudjelas, S. & De Poorter, M. (2000) 100 of the world's worst invasive alien species: a selection from the global invasive species database. Published by The Invasive Species Specialist Group (ISSG) a specialist group of the Species Survival Commission (SSC) of the World Conservation Union (IUCN), 12pp. First published as special lift-out in *Aliens*, 12.

McGeoch, M.A., Butchart, S.H.M., Spear, D., Marais, E., Kleynhans, E.J., Symes, A. et al. (2010) Global indicators of biological invasion: species numbers, biodiversity impact and policy responses. *Diversity and Distributions*, 16, 95–108.

Melbourne, B.A., Cornell, H.V., Davies, K.F., Dugaw, C.J., Elmendorf, S., Freestone, A.L. et al. (2007) Invasion in a heterogeneous world: resistance, coexistence or hostile takeover? *Ecology Letters*, 10, 77–94.

Mikissa, J.B., Jeffery, K., Fresneau, D. & Mercier, J.L. (2013) Impact of an invasive alien ant, *Wasmannia auropunctata* Roger., on a specialised plant-ant mutualism, *Barteria fistulosa* Mast. and *Tetraponera aethiops* F. Smith., in a Gabon forest. *Ecological Entomology*, 38, 580–584.

Moffett, M.W. (2012) Supercolonies of billions in an invasive ant: what is a society? *Behavioral Ecology*, 23, 925–933.

Noordijk, J. (2016) Leefwijze van *Tapinoma nigerrimum* (Hymenoptera: Formicidae), een nieuwe exotische mier in Nederland. *Entomologische Berichten*, 76, 86–93.

Palmer, T.M., Stanton, M.L. & Young, T.P. (2003) Competition and coexistence: exploring mechanisms that restrict and maintain diversity within mutualist guilds. *The American Naturalist*, 162, 63–79.

Palmer, T.M., Young, T.P. & Stanton, M.L. (2002) Burning bridges: priority effects and the persistence of a competitively subordinate acacia-ant in Laikipia, Kenya. *Oecologia*, 133, 372–379.

Passera, L. (1994) *Characteristics of tramp species*. In: Williams D (ed) Exotic Ants: biology, impact and control of introduced species. Boulder CO: Westview Press, pp. 23–43.

Pedersen, E.J., Miller, D.L., Simpson, G.L. & Ross, N. (2019) Hierarchical generalized additive models in ecology: an introduction with mgcv. *PeerJ*, 7, e6876.

Perales, K.M., Hansen, G.J.A., Hein, C.L., Mrnak, J.T., Roth, B.M., Walsh, J.R. et al. (2021) Spatial and temporal patterns in native and invasive crayfishes during a 19-year whole-lake invasive crayfish removal experiment. *Freshwater Biology*, 66, 2105–2117.

Pergl, J., Pyšek, P., Bacher, S., Essl, F., Genovesi, P., Harrower, C.A. et al. (2017) Troubling travellers: are ecologically harmful alien species associated with particular introduction pathways? *NeoBiota*, 32, 1–20.
*Tapinoma nigerrimum* complex (Hymenoptera: Formicidae) are related to sensitivity to urbanization. *Myrmecological News*, 32, 41–50.

Cleland, E.E., Esch, E. & McKinney, J. (2015) Priority effects vary with species identity and origin in an experiment varying the timing of seed arrival. *Oikos*, 124, 33–40.

Cordonnier, M., Blight, O., Angulo, E. & Courchamp, F. (2020) The native ant *Lasius niger* can limit the access to resources of the invasive Argentine ant. *Animals*, 10, 2451.

Cros, S., Cerdá, X. & Retana, J. (1997) Spatial and temporal variations in the activity patterns of Mediterranean ant communities. *Écoscience*, 4, 269–278.

Davidson, D.W. (1998) Resource discovery versus resource domination in ants: a functional mechanism for breaking the trade-off. *Ecological Entomology*, 23, 484–490.

de Soares, S.A. (2013) The role of competition in structuring ant communities: a review. *Oecologia Australis*, 17, 271–281.

Dekoninck, W., Parmentier, T. & Seifert, B. (2015) First records of a super-colonial species of the *Tapinoma nigerrimum* complex in Belgium (Hymenoptera: Formicidae). *Bulletin van de Koninklijke Belgische Vereniging voor Entomologie*, 151, 206–209.

Diagne, C., Leroy, B., Vaissière, A.-C., Gozlan, R.E., Roiz, D., Jarić, I. et al. (2021) High and rising economic costs of biological invasions worldwide. *Nature*, 592, 571–576.

Didham, R.K., Tylianakis, J.M., Hutchison, M.A., Ewers, R.M. & Gemmell, N.J. (2005) Are invasive species the drivers of ecological change? *Trends in Ecology & Evolution*, 20, 470–474.

Dijkstra, J.A., Westerman, E.L. & Harris, L.G. (2017) Elevated seasonal temperatures eliminate thermal barriers of reproduction of a dominant invasive species: a community state change for northern communities? *Diversity and Distributions*, 23, 1182–1192.

Fukami, T. (2015) Historical contingency in community assembly: integrating niches, species pools, and priority effects. *Annual Review of Ecology, Evolution, and Systematics*, 46, 1–23.

Gan, J.Y.S., Pathy, T.S., Yeo, D.Y.C., Lee, J.M. & Yek, S.H. (2022) The population dynamics of the yellow crazy ant *Anoplolepis gracilipes* (Hymenoptera: Formicidae) on a tropical Island in Malaysia. *Ecological Entomology*, 47, 339–346.

Gippet, J.M.W., George, L. & Bertelsmeier, C. (2022) Local coexistence of native and invasive ant species is associated with micro-spatial shifts in foraging activity. *Biological Invasions*, 24, 761–773.

Gouraud, C. & Kaufmann, B. (2022) Nouvelles observations des fourmis invasives du complexe des *Tapinoma* gr. *nigerrimum* dans le Massif armoricain (Hymenoptera: Formicidae). *Invertébrés Armoricains*, 23, 23–38.

Guo, F., Guénard, B., Economo, E.P., Deutsch, C.A. & Bonebrake, T.C. (2020) Activity niches outperform thermal physiological limits in predicting global ant distributions. *Journal of Biogeography*, 47, 829–842.

Helanterä, H. (2022) Supercolonies of ants (Hymenoptera: Formicidae): ecological patterns, behavioural processes and their implications for social evolution. *Myrmecological News*, 32, 1–22.

Hoffmann, B., Davis, P., Gott, K., Jennings, C., Joe, S., Krushelnycky, P. et al. (2011) Improving ant eradications: details of more successes, a global synthesis and recommendations. *Aliens*, 31, 16–23.

Hoffmann, B.D., Luque, G.M., Bellard, C., Holmes, N.D. & Donlan, C.J. (2016) Improving invasive ant eradication as a conservation tool: a review. *Biological Conservation*, 198, 37–49.

Holway, D.A., Lach, L., Suarez, A.V., Tsutsui, N.D. & Case, T.J. (2002) The causes and consequences of ant invasions. *Annual Review of Ecology and Systematics*, 33, 181–233.

Hulme, P.E. (2009) Trade, transport and trouble: managing invasive species pathways in an era of globalization. *Journal of Applied Ecology*, 46, 10–18.

INSEE. (2022). https://www.insee.fr/fr/statistiques/1405599?geo=DEP-34+COM-34172. Accessed on 13 May 2024.

Jung, J.-M., Lee, H.-S., Lee, J.-H., Jung, S. & Lee, W.-H. (2021) Development of a predictive model for soil temperature and its application to species distribution modeling of ant species in South Korea. *Ecological Informatics*, 61, 101220.

KimFung, C. & ChowYang, L. (2009) Influences of temperature, relative humidity and light intensity on the foraging activity of field populations of the longlegged ant, *Anoplolepis gracilipes* (Hymenoptera: Formicidae). *Sociobiology*, 54, 531–540.

Krushelnycky, P.D. & Gillespie, R.G. (2010) Correlates of vulnerability among arthropod species threatened by invasive ants. *Biodiversity and Conservation*, 19, 1971–1988.

Lei, Y., Zhou, Y., Lü, L. & He, Y. (2019) Rhythms in foraging behavior and expression patterns of the foraging gene in *Solenopsis invicta* (Hymenoptera: Formicidae) in relation to photoperiod. *Journal of Economic Entomology*, 112, 2923–2930.

Lenoir, A., Mercier, J.-L., Perdereau, E., Berville, L. & Galkowski, C. (2023) Sur l'expansion des fourmis envahissantes du genre *Tapinoma* en France (Hymenoptera: Formicidae). *Osmia*, 11, 1–10.

Lord, J.P. (2017) Impact of seawater temperature on growth and recruitment of invasive fouling species at the global scale. *Marine Ecology*, 38, e12404.

Lowe, S., Browne, M., Boudjelas, S. & De Poorter, M. (2000) 100 of the world's worst invasive alien species: a selection from the global invasive species database. Published by The Invasive Species Specialist Group (ISSG) a specialist group of the Species Survival Commission (SSC) of the World Conservation Union (IUCN), 12pp. First published as special lift-out in *Aliens*, 12.

McGeoch, M.A., Butchart, S.H.M., Spear, D., Marais, E., Kleynhans, E.J., Symes, A. et al. (2010) Global indicators of biological invasion: species numbers, biodiversity impact and policy responses. *Diversity and Distributions*, 16, 95–108.

Melbourne, B.A., Cornell, H.V., Davies, K.F., Dugaw, C.J., Elmendorf, S., Freestone, A.L. et al. (2007) Invasion in a heterogeneous world: resistance, coexistence or hostile takeover? *Ecology Letters*, 10, 77–94.

Mikissa, J.B., Jeffery, K., Fresneau, D. & Mercier, J.L. (2013) Impact of an invasive alien ant, *Wasmannia auropunctata* Roger., on a specialised plant-ant mutualism, *Barteria fistulosa* Mast. and *Tetraponera aethiops* F. Smith., in a Gabon forest. *Ecological Entomology*, 38, 580–584.

Moffett, M.W. (2012) Supercolonies of billions in an invasive ant: what is a society? *Behavioral Ecology*, 23, 925–933.

Noordijk, J. (2016) Leefwijze van *Tapinoma nigerrimum* (Hymenoptera: Formicidae), een nieuwe exotische mier in Nederland. *Entomologische Berichten*, 76, 86–93.

Palmer, T.M., Stanton, M.L. & Young, T.P. (2003) Competition and coexistence: exploring mechanisms that restrict and maintain diversity within mutualist guilds. *The American Naturalist*, 162, 63–79.

Palmer, T.M., Young, T.P. & Stanton, M.L. (2002) Burning bridges: priority effects and the persistence of a competitively subordinate acacia-ant in Laikipia, Kenya. *Oecologia*, 133, 372–379.

Passera, L. (1994) *Characteristics of tramp species*. In: Williams D (ed) Exotic Ants: biology, impact and control of introduced species. Boulder CO: Westview Press, pp. 23–43.

Pedersen, E.J., Miller, D.L., Simpson, G.L. & Ross, N. (2019) Hierarchical generalized additive models in ecology: an introduction with mgcv. *PeerJ*, 7, e6876.

Perales, K.M., Hansen, G.J.A., Hein, C.L., Mrnak, J.T., Roth, B.M., Walsh, J.R. et al. (2021) Spatial and temporal patterns in native and invasive crayfishes during a 19-year whole-lake invasive crayfish removal experiment. *Freshwater Biology*, 66, 2105–2117.

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Pyšek, P., Bacher, S., Kühn, I., Novoa, A., Catford, J.A., Hulme, P.E. et al. (2020) MAcroecological framework for invasive aliens (MAFIA): disentangling large-scale context dependence in biological invasions. *NeoBiota*, 62, 407-461.

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Ribas, C.R. & Schoereder, J.H. (2002) Are all ant mosaics caused by compe-tition? *Oecologia*, 131, 606-611.

Roeder, K.A., Roeder, D.V. & Kaspari, M. (2018) The role of temperature in competition and persistence of an invaded ant assemblage. *Ecological Entomology*, 43, 774-781.

RStudio Core Team. (2020) RStudio: integrated development environment for R. Boston, MA: RStudio, PBC.

Ruano, F., Tinaaut, A. & Soler, A.J.J. (2000) High surface temperatures select for individual foraging in ants. *Behavioral Ecology*, 11, 396-404.

Ruland, F. & Jeschke, J.M. (2020) How biological invasions affect animal behaviour: a global, cross-taxonomic analysis. *Journal of Animal Ecol-ogy*, 89, 2531-2541.

Seebens, H., Blackburn, T.M., Dyer, E.E., Genovesi, P., Hulme, P.E., Jeschke, J.M. et al. (2017) No saturation in the accumulation of alien species worldwide. *Nature Communications*, 8, 14435.

Seifert, B., D'Eustacchio, D., Kaufmann, B., Centorame, M., Lorite, P. & Modica, M.V. (2017) Four species within the supercolonial ants of the *Tapinoma nigerrimum* complex revealed by integrative taxonomy (Hymenoptera: Formicidae). *Myrmecological News*, 24, 123-144.

Shi, J., Luo, Y.-Q., Zhou, F. & He, P. (2010) The relationship between inva-sive alien species and main climatic zones. *Biodiversity and Conserva-tion*, 19, 2485-2500.

Simberloff, D., Martin, J.-L., Genovesi, P., Maris, V., Wardle, D.A., Aronson, J. et al. (2013) Impacts of biological invasions: what's what and the way forward. *Trends in Ecology & Evolution*, 28, 58-66.

Simpson, G.L. (2018) Modelling palaeoecological time series using generalised additive models. *Frontiers in Ecology and Evolution*, 6, 149.

Tang, X., Yuan, Y., Liu, X. & Zhang, J. (2021) Potential range expan-sion and niche shift of the invasive *Hyphantria cunea* between native and invasive countries. *Ecological Entomology*, 46, 910-925.

Valladares, F., Bastias, C.C., Godoy, O., Granda, E. & Escudero, A. (2015) Species coexistence in a changing world. *Frontiers in Plant Science*, 6, 866.

Warren, R.J., Il & Chick, L. (2013) Upward ant distribution shift corre-sponds with minimum, not maximum, temperature tolerance. *Global Change Biology*, 19, 2082-2088.

Weidlich, E.W.A., Nelson, C.R., Maron, J.L., Callaway, R.M., Delory, B.M. & Temperton, V.M. (2021) Priority effects and ecological restoration. *Restoration Ecology*, 29, e13317.

Wetterer, J.K., Espadaler, X., Wetterer, A.L., Aguin-Pombo, D. & Franquinho-Aguia, A.M. (2006) Long-term impact of exotic ants on the native ants of Madeira. *Ecological Entomology*, 31, 358-368.

Wetterer, J.K. & Moore, J.A. (2005) Red imported fire ants (Hymenoptera: Formicidae) at gopher tortoise (Testudines: Testudinidae) burrows. *The Florida Entomologist*, 88, 349-354.

Williamson, M. & Fitter, A. (1996) The varying success of invaders. *Ecology*, 77, 1661-1666.

Wood, S.N. (2003) Thin plate regression splines. *Journal of the Royal Statistical Society, Series B: Statistical Methodology*, 65, 95-114.

Wood, S.N. (2006) *Generalized additive models: an introduction with R*. New York, NY: Chapman and Hall/CRC.

Wood, S.N. (2011) Fast stable restricted maximum likelihood and marginal likelihood estimation of semiparametric generalized linear models. *Journal of the Royal Statistical Society, Series B: Statistical Methodology*, 73, 3-36.

Wood, S.N. (2017) *Generalized additive models: an introduction with R*, 2nd edition, Boca Raton, FL: Chapman and Hall/CRC.

Wyatt, S.A. (2004) Ant (*Formicidae*) diversity, species richness, and abundance in response to different food sources. *Tropical Ecology and Conservation* [Monteverde Institute]. https://digitalcommons.usf.edu/tropical_ecology/130

Young, S.L., Clements, D.R. & DiTommaso, A. (2017) Climate dynamics, invader fitness, and ecosystem resistance in an invasion-factor framework. *Invasive Plant Science and Management*, 10, 215-231.

## SUPPORTING INFORMATION

Additional supporting information can be found online in the Support-
ing Information section at the end of this article.

<caption>**Figure S1.** Mean worker abundance (±SE) depending on temperature categories for (A) *T. darioi* (orange) and *T. magnum* (green) (B) *Tapinoma* (red) species and native species (blue).</caption>

**Table S1.** Characteristics of sampled sites.

**Table S2.** Raw data related to activity measurements. «temp air beginning» and «temp air end» indicate the air temperature at the beginning and the end of each measurements; «temp ground middle», «temp ground right» and «temp ground left» represent the temperature measured 10 cm below the ground at the central point and on the two sides respectively; «type of bait» can either be honey (H), peanut butter (P) or tuna rillettes (T).

**How to cite this article:** Destour, G., Blatrix, R., Caries, A., Genevet, J., Kaufmann, B., Lecoq, B. et al. (2024) Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area. *Ecological Entomology*, 1-9. Available from: https://doi.org/10.1111/een.13339
Pol, R. & de Casenave, J.L. (2004) Activity patterns of harvester ants *Pogonomyrmex pronotalis* and *Pogonomyrmex rastratus* in the Central Monte Desert, Argentina. *Journal of Insect Behavior*, 17, 647-661.

Pyšek, P., Bacher, S., Kühn, I., Novoa, A., Catford, J.A., Hulme, P.E. et al. (2020) MAcroecological framework for invasive aliens (MAFIA): disentangling large-scale context dependence in biological invasions. *NeoBiota*, 62, 407–461.

R Core Team. (2022) R: a language and environment for sta-tistical comput-ing. Vienna: R Foundation for Statistical Computing.

Rahardjo, B.T., Muhammad, F.N., Setiawan, Y., Febryadi, A., Ihsan, M., Wibowo, D. et al. (2023) Ant preference for different types of bait at sugarcane plantations in East Java, Indonesia. *Biodiversitas Journal of Biological Diversity*, 24, 2099–2106.

Ribas, C.R. & Schoereder, J.H. (2002) Are all ant mosaics caused by compe-tition? *Oecologia*, 131, 606–611.

Roeder, K.A., Roeder, D.V. & Kaspari, M. (2018) The role of temperature in competition and persistence of an invaded ant assemblage. *Ecological Entomology*, 43, 774–781.

RStudio Core Team. (2020) RStudio: integrated development environment for R. Boston, MA: RStudio, PBC.

Ruano, F., Tinaaut, A. & Soler, A.J.J. (2000) High surface temperatures select for individual foraging in ants. *Behavioral Ecology*, 11, 396–404.

Ruland, F. & Jeschke, J.M. (2020) How biological invasions affect animal behaviour: a global, cross-taxonomic analysis. *Journal of Animal Ecol-ogy*, 89, 2531–2541.

Seebens, H., Blackburn, T.M., Dyer, E.E., Genovesi, P., Hulme, P.E., Jeschke, J.M. et al. (2017) No saturation in the accumulation of alien species worldwide. *Nature Communications*, 8, 14435.

Seifert, B., D'Eustacchio, D., Kaufmann, B., Centorame, M., Lorite, P. & Modica, M.V. (2017) Four species within the supercolonial ants of the *Tapinoma nigerrimum* complex revealed by integrative taxonomy (Hymenoptera: Formicidae). *Myrmecological News*, 24, 123-144.

Shi, J., Luo, Y.-Q., Zhou, F. & He, P. (2010) The relationship between inva-sive alien species and main climatic zones. *Biodiversity and Conserva-tion*, 19, 2485-2500.

Simberloff, D., Martin, J.-L., Genovesi, P., Maris, V., Wardle, D.A., Aronson, J. et al. (2013) Impacts of biological invasions: what's what and the way forward. *Trends in Ecology & Evolution*, 28, 58–66.

Simpson, G.L. (2018) Modelling palaeoecological time series using generalised additive models. *Frontiers in Ecology and Evolution*, 6, 149.

Tang, X., Yuan, Y., Liu, X. & Zhang, J. (2021) Potential range expan-sion and niche shift of the invasive *Hyphantria cunea* between native and invasive countries. *Ecological Entomology*, 46, 910-925.

Valladares, F., Bastias, C.C., Godoy, O., Granda, E. & Escudero, A. (2015) Species coexistence in a changing world. *Frontiers in Plant Science*, 6, 866.

Warren, R.J., Il & Chick, L. (2013) Upward ant distribution shift corre-sponds with minimum, not maximum, temperature tolerance. *Global Change Biology*, 19, 2082–2088.

Weidlich, E.W.A., Nelson, C.R., Maron, J.L., Callaway, R.M., Delory, B.M. & Temperton, V.M. (2021) Priority effects and ecological restoration. *Restoration Ecology*, 29, e13317.

Wetterer, J.K., Espadaler, X., Wetterer, A.L., Aguin-Pombo, D. & Franquinho-Aguia, A.M. (2006) Long-term impact of exotic ants on the native ants of Madeira. *Ecological Entomology*, 31, 358–368.

Wetterer, J.K. & Moore, J.A. (2005) Red imported fire ants (Hymenoptera: Formicidae) at gopher tortoise (Testudines: Testudinidae) burrows. *The Florida Entomologist*, 88, 349–354.

Williamson, M. & Fitter, A. (1996) The varying success of invaders. *Ecology*, 77, 1661–1666.

Wood, S.N. (2003) Thin plate regression splines. *Journal of the Royal Statistical Society, Series B: Statistical Methodology*, 65, 95–114.

Wood, S.N. (2006) Generalized additive models: an introduction with R. New York, NY: Chapman and Hall/CRC.

Wood, S.N. (2011) Fast stable restricted maximum likelihood and marginal likelihood estimation of semiparametric generalized linear models. *Journal of the Royal Statistical Society, Series B: Statistical Methodology*, 73, 3–36.

Wood, S.N. (2017) Generalized additive models: an introduction with R, 2nd edition, Boca Raton, FL: Chapman and Hall/CRC.

Wyatt, S.A. (2004) Ant (Formicidae) diversity, species richness, and abundance in response to different food sources. *Tropical Ecology and Conservation* [Monteverde Institute]. https://digitalcommons.usf.edu/tropical_ecology/130

Young, S.L., Clements, D.R. & DiTommaso, A. (2017) Climate dynamics, invader fitness, and ecosystem resistance in an invasion-factor framework. *Invasive Plant Science and Management*, 10, 215–231.

SUPPORTING INFORMATION

Additional supporting information can be found online in the Support-
ing Information section at the end of this article.

<caption>**Figure S1.** Mean worker abundance (±SE) depending on temperature categories for (A) *T. darioi* (orange) and *T. magnum* (green) (B) *Tapinoma* (red) species and native species (blue).</caption>

**Table S1.** Characteristics of sampled sites.

**Table S2.** Raw data related to activity measurements. «temp air beginning» and «temp air end» indicate the air temperature at the beginning and the end of each measurements; «temp ground middle», «temp ground right» and «temp ground left» represent the temperature measured 10 cm below the ground at the central point and on the two sides respectively; «type of bait» can either be honey (H), peanut butter (P) or tuna rillettes (T).

How to cite this article: Destour, G., Blatrix, R., Caries, A., Genevet, J., Kaufmann, B., Lecoq, B. et al. (2024) Impact of temperature and season on ant foraging activity of two invasive species in a Mediterranean urban area. *Ecological Entomology*, 49(5), 673–681. Available from: https://doi.org/
10.1111/een.13339