Mixed responses of ant assemblages to Carpobrotus spp. and Rattus rattus eradications on a small Mediterranean island

doi:10.21203/rs.3.rs-3551654/v1

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Mixed responses of ant assemblages to Carpobrotus spp. and Rattus rattus eradications on a small Mediterranean island

https://doi.org/10.21203/rs.3.rs-3551654/v1

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Eradication of invasive alien species (IAS) is often proposed to restore invaded ecosystems, with information on subsequent ecosystem recovery key to conservation policies. Although ants perform major ecological functions in the ecosystem, their response to IAS eradication has received relatively little monitoring. This study investigated ant response to Carpobrotus spp. (iceplant) and R. rattus (black rat) eradications on the small Mediterranean island of Bagaud (Var, France). Ant assemblages were monitored over a ten-year period, including two years before eradications, at six different sites, two invaded by Carpobrotus spp., two under high R. rattus pressure, and two control sites. We found inter-annual variations in ant assemblages but no before-after eradication trend at both control and rat eradication sites. However, there was a clear increase in ant species richness and abundance after the Carpobrotus eradication. A trend toward more homogeneous assemblages was observed, with an enhanced core of common Mediterranean ant species. Xerophilous and thermophilous species such as Pheidole pallidula, Messor bouvieri, and Plagiolepis pygmaea were more abundant, consistent with the return of native vegetation communities and possibly warmer and dryer microclimatic conditions. The trend was even stronger on the denser and thicker Carpobrotus eradication patch. Increased ant foraging activity is a good sign of restoration success, influencing different functions of the ecosystem, such as seed dispersal or nutrient cycling.

Formicidae

Restoration

Exotic

Monitoring

Community Recovery

Islands represent a major challenge for biodiversity conservation, hosting high levels of endemism that exceed those of mainland regions - for example, eight times higher for vertebrates (Kier et al. 2009). Yet much of this biodiversity is being lost, threatened by multiple factors including biological invasions (Fernández-Palacios et al. 2021). Compared to mainland regions, islands suffer higher invasion rates (Dawson et al. 2017), with invasive alien species (IAS) identified as one of the main causes of species extinctions (Pyšek et al. 2020). Invasions often cause changes in community structure and modify abiotic characteristics of ecosystems (Strayer et al. 2006; Pyšek et al. 2012). Insular ecosystems are remarkably vulnerable due to their small size, small populations and vacant niches without competitors or predators (Fernández-Palacios et al. 2021; Russell et al. 2017).

Considered as a hotspot of biodiversity, the Mediterranean basin and its numerous islands (>10,000) are highly threatened by invasions (Myers et al. 2000; Hulme et al. 2008) and could be the ecosystems most impacted by IAS by 2100 (Sala et al. 2000). Many islands are now faced with numerous interacting invasive species, such as plants and seed dispersers, which pose a severe threat to the native biodiversity (Courchamp et al. 2011; Russell and Kaiser-Bunbury 2019).

Such invasive meltdown was described for two invasive taxa, the black rat (Rattus rattus) and iceplants (Carpobrotus spp., Aizoceae), which promote each other (Bourgeois et al. 2005). The black rat (Rattus rattus) is considered as one of the 15 most prevalent IAS on islands (Russell et al. 2017). It has colonized most of the Mediterranean islands over the last 2,000 years (Ruffino et al. 2009), with direct effects on biodiversity, through predation. The black rat is widely known to prey on seabird eggs and chicks, and on large-bodied invertebrates, such as beetles, spiders, gastropods (Caut et al. 2009; Nascimento et al. 2019). It also has indirect effects by dispersing invasive plant seeds, including Carpobrotus species (Bourgeois et al. 2005; Shiels 2011).

Also known as iceplants or hottentot figs, Carpobrotus species were introduced to Europe in the seventeenth century from South Africa, Australia or Chile for ornamental and soil stabilization purposes (Campoy et al. 2018). Carpobrotus acinaciformis, Carpobrotus edulis and their hybrids (Carpobrotus aff. acinaciformis) have now colonized a large portion of the western Mediterranean coast (Suehs et al. 2004; Campoy et al. 2018). They establish in various coastal environments, from sandy dunes to rocky cliffs, forming dense monospecific mats with an abundant litter (Campoy et al. 2018). The succulent species of Carpobrotus (Aizoceae) are a major threat to native biodiversity in Mediterranean coastal ecosystems (Hulme et al. 2008; Carboneras et al. 2018). They shift soil chemical parameters (Novoa et al. 2013a; Vieites-Blanco and González-Prieto 2018), decrease plant diversity (Badalamenti et al. 2016), and reduce invertebrate abundance and modify community composition (Rodríguez et al. 2020).

In order to mitigate the adverse impacts of IAS on biodiversity, control or eradication operations are carried out, prioritizing the most prevalent IAS, including the black rat and the iceplant (Carboneras et al. 2018). In the Mediterranean Basin, controls and eradications have generally led to a positive response in terms of biodiversity (Jones et al. 2016). Carpobrotus eradications have been shown to positively impact native plants (Lazzaro et al. 2020; Buisson et al. 2021). Rattus rattus eradications have benefited reptiles (Krebs et al. 2015) and breeding bird populations (Howald et al. 2010; Bourgeois et al. 2013). However, while several studies investigated invertebrate recovery after rat eradication (St Clair et al. 2011; Ruscoe et al. 2013) and Carpobrotus removal (St Clair et al. 2011; Braschi et al. 2021b, a), these did not include ants.

Ubiquitous and abundant, ants play a key role in ecosystem functioning, in soil structure, as predators, or as promoters of plant dispersion (Del Toro et al. 2012; De Almeida et al. 2020). They are sensitive and respond quickly to environmental change, which makes them good bioindicators (Majer 1983; Andersen et al. 2002). Habitat structure, microclimate, resource availability and competitive interactions are indirect factors responsible for ant community modifications (Andersen 2019). The decrease in ant diversity correlated with habitat simplification and loss of resources like seeds points to the need to monitor ant communities, a crucial component of the ecosystem, in restoration projects (Underwood and Fisher 2006). Very few studies focus on ant response to plant or mammal IAS eradication (Parr et al. 2010; Osunkoya et al. 2011; Lenda et al. 2013), although ant monitoring is a relevant way of assessing ecosystem change (Kaspari and Majer 2000).

The aim of this study was to explore ant responses to Carpobrotus spp. and R. rattus eradications on the small Mediterranean island of Bagaud (Var, France). Ant assemblages were compared over a ten-year monitoring period, including two years before eradications, at six different sites on the island: two sites invaded by Carpobrotus spp., two sites under high R. rattus pressure, and the last two as controls. We hypothesized that : (i) Ant species richness would increase after the IAS eradication at the four invaded sites and remain stable at the control sites; (ii) Ant abundance would increase at the eradication sites and remain stable at the control sites; and (iii) The composition of ant assemblages would change after the IAS eradication at the four invaded sites and remain stable at the control sites; (iv) The magnitude of changes in ant assemblage structure and composition would be greater after the eradication of Carpobrotus spp. than of R. rattus, as Carpobrotus spp. would induce more habitat modifications.

1.1 Study area

Bagaud island is a small uninhabited Mediterranean island of 58 ha located in the Hyères archipelago (Var, France), which is part of the Port-Cros National Park. It is a strict biological reserve with restricted access since 2007. The continent is 7.5km away and the main nearest island, Port-Cros, is 1 km away. The archipelago has a Mediterranean climate with annual mean temperatures ranging between 13.5 and 19.5°C and an average 621.5 mm of precipitations per year (1991-2020, Ile du Levant weather station, Infoclimat). Vegetation on Bagaud island is typical of Mediterranean regions, with dominant matorral and halophilous plants on the rocky shores. In addition to its rich native plant communities, the invasive plant species Carpobrotus edulis and Carpobrotus aff. acinaciformis were introduced in the nineteenth century by military engineers. Intended to stabilize the soil around the island’s forts, these species covered 2-ha of the island at their worst. The only non-flying mammal of the island, the black rat (R. rattus), was also introduced, probably during the Roman period (Ruffino et al. 2015). It preys on Carpobrotus spp. fruits as well as potentially on various seeds, beetles, geckos, and seabird eggs and chicks (Ruffino et al. 2015). Eradications of Carpobotus spp. and R. rattus were carried out on the island in late 2011 and 2012.

1.2 Study sites

We monitored the temporal dynamics of ant communities via three treatments: (1) Carpobrotus spp. removal (hereafter CA), (2) Rattus rattus removal (hereafter RR) and (3) control sites with no eradication action (hereafter CO) (Fig.1). Limited availability of sites led us to select two sites per treatment only. The two Carpobrotus spp. removal sites were selected for size, large enough to set up transects of 10 pitfall traps (see section 1.3). The two Rattus rattus removal sites were selected as having the highest rate of Rattus rattus captures and being accessible for arthropod sampling.

Carpobrotus spp. was eradicated from study sites CA1 and CA2 between November 2011 and February 2012. Plants were uprooted and left to compost on the site to avoid further dissemination (Chenot et al. 2018). Control operations have been conducted every year since then to prevent the recolonization of the area from the seed bank (Ruffino et al. 2015; Buisson et al. 2021).

Rats were trapped from September 2011 to June 2012 at RR1 and RR2 sites. Rat eradication was carried out first via trapping sessions and later supplemented by rodenticide (Ruffino et al. 2015). Twenty permanent bait stations were set up along the coast under biosecurity measures (Ruffino et al. 2015). Traces on baits were spotted again in 2013, and while biosecurity increased, rat traces became significantly more numerous in 2017, fluctuating thereafter (Braschi 2021).

Our two control sites were selected as presenting the lowest rates of rat captures, i.e. below 37 individuals / hectare (Ruffino et al. 2015), and no Carpobrotus ever reported. CO1 and CO2 sites represent the island’s two main vegetation types (low and high matorrals). A habitat description for each study site (CA1 and CA2, RR1 and RR2, CO1 and CO2) is available in Table 1.

1.3 Sampling design

Ants were collected using pitfall traps, a method commonly used to sample ground-dwelling arthropods and which has proved efficient in assessing ant species richness and spatial patterns (Steiner et al. 2005; Salata et al. 2020). The two first years of sampling, 2010 and 2011, were pre-eradication, while the following four surveys were carried out post-eradication, i.e. in 2013, 2015, 2017, and 2019. Each year, 10 pitfall traps (5 cm diameter and 10 cm depth), half-filled with ethylene-glycol, were installed 5 meters apart on a 45-meter transect on each site. They were collected every three weeks, four times from late April to late June / early July, and again collected between mid-September and mid-October, a total of four sampling periods. Ants were counted and identified to species level or to species-group level, following Lebas et al. (2016).

1.4 Data analysis

Statistical analyses were performed using R version 4.2.2 (2022-10-31). The temporal dynamics of ant communities was studied for each site independently. Sites were not statistically compared with each other due to the limited number of true site replicates (see section 1.2) (as in Braschi et al. 2021b, a). First, changes in species richness and total abundance of ants over time were assessed. Each trap’s species richness was calculated. Raw ant counts were not used because these depend on species-specific social behavior (Steiner et al. 2005) and do not reflect the natural unit of ant diversity, the individual nest (Gotelli et al. 2011). Thus, the abundance of each species was transformed on a 7-score scale (Hoffmann et al. 2000): 1 = 1 ant; 2 = 2-5 ants; 3 = 6-10 ants; 4 = 11-20 ants; 5 = 21-50 ants; 6 = 51-100; and 7 > 100 ants. Total abundance was obtained by summing the species scores for each trap. Each site was studied independently, using a generalized linear mixed model from the glmmTMB R package (Brooks et al., 2017), with a negative binomial distribution (logit link) with either species richness or total abundance as response variables. The year was set as a fixed factor and the period of sampling and the trap number were set as random factors. Multiple comparisons between years were performed using the Sidak method.

The assemblage compositions of all sites were compared over the six years of survey using the non-metric multidimensional scaling (NMDS) approach with the vegan package (Oksanen et al. 2022). For this analysis, abundance classes were averaged over the four sampling periods of each year for each trap. Assemblages were represented using a Bray-Curtis dissimilarity matrix. Overall significant difference between assemblages was assessed using PERMANOVA, followed by pairwise adonis tests (Martinez Arbizu 2020). Further individual species abundance analyses were conducted for the sites showing significant dissimilarity. SIMPER analysis (Clarke 1993) was used to identify species contributing up to 75% of the assemblage dissimilarity between the pre-eradication grouped years and the post-eradication grouped years. Their individual abundance response to the year factor was assessed using glmmTMB with the same features as the total abundance models.

A total of 35,414 specimens, 24 species, and 13 genera were identified during the study (Appendix A). The majority of the individuals belonged to the Myrmicinae family (25,944 individuals), followed by Formicinae (9,491 individuals), and a few Ponerinae were also counted (28 individuals). CO2 had the most species, 18, and CA2 had the least, 11 species. Pheidole pallidula was the most abundant and frequently found species. Statistical results of the fitted models are available in Appendix B.

2.1 Species richness and total abundance

RR2 had the overall highest mean species richness, with an average of 4.90 ± 1.98 species caught per trap, against 2.38 ± 1.25 for RR1, the overall lowest value. Species richness per trap remained unchanged during the pre-eradication surveys in 2010 and 2011 at all the studied sites. The two Carpobrotus removal sites showed a significant increase in species richness per trap after the eradication (Fig 1). In 2013, CA2 species richness was not significantly different from either the pre-eradication years or from all the post-eradication years. Only in years 2015 to 2019 was species richness significantly higher than in 2010. At CA1, the species richness of 2013 was significantly higher than that of 2010. Years 2015 and 2017 showed significantly higher values compared to the pre-eradication years, and values in 2019 were significantly higher than all the other years, reaching an average of 5.06 species per trap. RR2 mean species richness per trap changed slightly over time, with significantly higher values in 2019 than in 2013 and 2017 (Fig 2). The other sites did not show any significant differences between years.

RR2 also had the overall highest mean abundance, an average abundance of 10.55 ± 5.18, against the overall lowest average of 4.87 ± 2.91 per trap at CO2 (Fig 3). Total ant abundance per trap increased significantly after the eradication at CA1 (Fig 3). Abundance levels in 2010 and 2011 were significantly lower than in any other year of sampling. They were up to 5 times lower than in 2019, when total abundance reached 11.75 ± 4.13. The mean total abundance of 2019 traps was also significantly higher than in 2013 and 2015. CA2 also showed a slight increase in total ant abundance over the years, 2010 mean total abundance being significantly lower than in the other years. It then significantly increased in 2011 and 2013 and again in 2015 and 2019. At RR1, RR2 and CO1, abundances fluctuated, with no particular trend (Fig 3), although they were significantly higher in 2019 compared to 2010 at both RR1 and RR2. Ant abundances did not change over the years at CO2.

2.2 Assemblage composition

The assemblages of CA1, CA2, RR2 and CO2 showed a significant percentage of variance according to year grouping. With an R² of 32%, CA1 assemblages differed the most across years, followed by CA2 with 19% and RR2 and CO2 with respectively 14% and 15%. At the two Carpobrotus eradication sites, 2010 and 2011 assemblages were the most heterogeneous (Fig 4). The CA1 assemblages of 2010 and 2011 were similar to each other but both different from all the subsequent years (2013-2019). The assemblage composition of 2019, the most homogeneous, was significantly different from all the previous years. Assemblages shifted over the years, from heterogeneous pre-eradication assemblages driven by Hypoponera eduardi to more homogeneous ones with a strong species core composed of Pheidole pallidula, Plagiolepis pygmaea, Crematogaster scutellaris, and Crematogaster sordidula. This trend was weaker at CA2, where polygons overlapped much more. The 2010 assemblage composition differed from all the subsequent years and the 2011 composition was only significantly different from 2019. The pairwise analysis also showed significant differences between 2013, 2015, and 2019. Like CA1, the CA2 assemblages became more homogeneous over the years, except for 2017, where the polygon was larger than in 2015. At RR2, the assemblage composition of 2010 was significantly different from all the other years except 2011. The assemblage of 2013 only differed from 2010 and 2017. With the largest polygon, the 2015 assemblage composition was the most heterogeneous (Fig 4). At CO2, the assemblages of 2010 and 2011 were significantly different from 2017 and 2019. At RR2 and CO2, polygons overlapped much more than at the Carpobrotus eradication sites (Fig 4).

2.3 Individual species abundance

At the CA1 site, three species contributed to 75% of the pre-/post-eradication assemblage dissimilarity: P. pallidula, Messor bouvieri, and P. pygmaea. For these three species, individual mean abundance increased after the eradication (Fig 5). This trend was stronger for P. pallidula and P. pygmaea, both showing abundance levels close to zero before Carpobrotus eradication, significantly lower than in the subsequent years. Their abundance steadily increased, reaching 4.12 ± 1.18 and 2.22 ± 1.48, respectively, in 2019. At CA2, the same three species contributed to most of the dissimilarity between pre- and post-eradication years. While P. pygmaea abundance did not change over the years, P. pallidula began to be significantly more abundant from 2015 on and M. bouvieri from 2017 on (Fig 5). At RR2, 75% of the pre-/post-eradication dissimilarity was attributed to six species: P. pallidula, Lasius gr. grandis, M. bouvieri, C. scutellaris, P. pygmaea, and Camponotus sylvaticus. Unlike the Carpobrotus eradication sites, no before-after trend was observed (Fig 5). The abundance level of P. pallidula was significantly lower in 2013 compared to 2010 and 2011. Lasius gr. grandis was significantly more abundant in 2019 compared to 2013 and 2017. Messor bouvieri abundance was significantly lower in 2015 and 2017 compared to 2010. In contrast, P. pygmaea was less abundant in 2010 compared to all subsequent years except 2013, while C. sylvaticus abundance was higher in 2013 than in 2010, 2011, and 2017. Concerning CO2, among the three main species contributing to pre-/post-eradication dissimilarity, only C. sordidula abundance was significantly influenced by year (Fig 6). It was significantly less abundant in 2010 and 2011 compared to the subsequent years.

The objective of the study was to assess changes in ant assemblage structure and composition after simultaneous eradication of Carpobrotus and R. rattus on a small Mediterranean island. As we expected, the temporal dynamics of ant assemblages showed a stronger trend toward pre-/post-eradication dissimilarity at Carpobrotus eradication sites (CA1 and CA2) than at rat eradication sites (RR1 and RR2).

At the control sites (CO1 et CO2), assemblage structures and compositions either remained stable or changed slightly, usually without showing any particular trend. At both sites, species richness remained stable and assemblage composition showed a strong overlap over the years. Crematogaster sordidula abundance was an exception, increasing for unknown reasons after 2011 at CO2.

Buisson et al. (2021) showed that the plant communities of Bagaud island remained relatively stable outside the Carpobrotus removal sites. The absence of habitat modifications or perturbations is consistent with relative stability in control assemblages. However, the slight changes in assemblage structure and composition observed across the years reflect a change in ant foraging activity, affecting trapping success (Steiner et al. 2005). These changes could be explained by climate variations (Pelini et al. 2014; Resasco et al. 2014) and/or by demographically volatile species (Samways 1990). The Mediterranean climate is known for the strong inter-annual variability of temperatures and precipitations (Deitch et al. 2017), and these parameters can strongly influence ant foraging activity (Whitford and Ettershank 1975; Levings 1983; Lasmar et al. 2021). Increased temperature can positively influence foraging activity, increasing ant abundance and diversity, with responses varying according to the thermal tolerance of the species (Stuble et al. 2013). In Mediterranean open vegetation, temperature is the main driver of changes in ground ant communities (Retana and Cerdá 2000). Apart from climatic variability, resource availability can also be a strong driver of demographic changes and foraging behavior in ants (Bernstein 1979; Davidson et al. 1985; Traniello 1989). Moreover, some species show changes in their nest distributions from year to year and highly flexible behavior (Gordon 1991).

Our sampling is composed of common Mediterranean heat-tolerant species that could benefit from open and dry habitats to increase their activity (Cros et al. 2016). According to the Porquerolles weather station, 2011 was the wettest and one of the coldest years of our study (Table 2). Xerophilous and thermophilous species could have reduced their foraging activity because of non-favorable climatic conditions, reducing their probability of being trapped, and explaining the slightly lower overall ant abundance that year. Nevertheless, the low abundance of Crematogaster sordidula in both 2010 and 2011 compared to the rest of the sampling is surprising, as the climate differed between these two years, 2011 being warmer and much dryer. Yet Crematogaster sordidula mean abundance did not exceed an average of one worker caught per trap. As its most common habitats are open and well exposed (Retana et al. 2015), workers could occasionally visit the high matorral site to forage from the nearby open areas which represent a favorable nesting habitat. A new nest closer to the sampling site could explain the higher worker abundance after 2011.

Similarly, no before-after pattern was observed at the rat eradication sites. Overall, 2019 had higher levels of abundance and species richness, and assemblages were more homogeneous. Individual responses were mixed, depending on species identity and sites. Although eradications initially decreased rat densities on Bagaud island, the population density rose again in 2015 and 2017. The lack of response from ants to rat eradication in the first years is consistent with previous findings that ants are not commonly a prey of Rattus rattus (Courchamp et al. 2011; Riofrío-Lazo and Páez-Rosas 2015). Moreover, ant abundance and species richness were already at similar or higher levels than those of the control sites. Other studies on rat eradication or exclusion failed to show any response in ant abundance (Sinclair et al. 2005; Vergara et al. 2021). One of them even showed a decrease in ant abundance, attributed to either the rodenticide or to ant behavioral change in response to higher resource availability (Rate 2009).

Variation in ant abundances and ant species richness at the rat eradication sites could, once again, be correlated with climatic variability. A warmer and dryer climate was recorded in 2019, especially compared to 2010 and 2013. Concerning individual species abundances, Lasius gr. grandis and Plagiolepis pygmaea, which are both thermophilous Mediterranean species, showed abundance variations consistent with higher activity under a warmer climate (Majeed et al. 2021; Santos et al. 2022). However, no particular trend was observed in Pheidole pallidula and Camponotus sylvaticus abundance variations. Abundance variability in such dominant and highly competitive species (Retana et al. 2015) has already been observed, with worker numbers multiplying as resource availability increases (Samways 1990). Similarly, the granivorous ant Messor bouvieri is highly dependent on seed availability (Willott et al. 2000), which can be influenced by vegetation dynamics and climate variation through complex interactions (Carmona et al. 2015).

The fact that inter-annual variations without a before-after trend were observed at the control and rat-invaded sites indicates that the changes observed in ant assemblages at Carpobrotus spp.-invaded sites can be attributed to the Carpobrotus eradication. Indeed, while the pool of species found at the CA sites did not change after eradication, we observed a shift toward more homogeneous assemblages. Carpobrotus eradication promoted an abundant species core composed of common xerophilous Mediterranean ant species. This trend was stronger at the Carpobrotus eradication site located on the cliff top and surrounded by low matorral (CA1), where we observed an overall steady increase in ant abundances.

These changes in the ant assemblages can be explained by modification of the ecosystem after eradication. As a succulent and with its ability to create dense litter and mats, Carpobrotus induces changes at the soil surface: it reduces evapotranspiration, incident solar radiation, and thus temperature (Molinari et al., 2007; Novoa et al., 2013). Native vegetation quickly recovered after eradication, inducing changes in habitats, from poor and homogeneous dense mats of Carpobrotus to higher plant species richness, a decrease in litter cover, and fluctuation of bare ground cover with seasons and years (Buisson et al. 2021). At a small scale, the return of native vegetation added habitat structure and resource diversity, and allowed a return to the warmer and dryer microclimate that benefits most Mediterranean ants.

Ants are sensitive to habitat disturbance through the indirect effects of change in habitat structure, microclimate, resource availability, and competitive interactions (Andersen 2019). Ant species richness and abundance are closely correlated to vegetation variables, such as vegetation cover and strata (Retana and Cerdá 2000; Lassau and Hochuli 2004). Diversified plant communities promote more heterogeneity in the vegetation structure, with varying plant heights and microhabitats (Buisson et al. 2021), which has already been shown to promote ant species diversity in various ecosystems (Vasconcelos et al. 2008; Hill et al. 2008; Martello et al. 2018).

The dynamics of the resource use of many ant species is strongly influenced by microclimatic conditions (Cerdá et al. 1998), in particular by fine-scale temperature changes (Stuble et al. 2013) and by microhabitat formation (Luque and Reyes López 2007). The diversification of vegetation also influences nest site availability (Herbers 1989) and temporal spread of certain food resources, such as nectar or seeds, contrasting with the few weeks of Carpobrotus flowering during spring (Campoy et al. 2018). The increased post-eradication abundance of ant species with a specialized diet, such as Plagiolepis pygmaea that feed on liquid food and Messor bouvieri that feed on seeds (Arnan et al. 2014), confirm the benefits of habitat diversification. However, removing the abundant litter produced by Carpobrotus might be detrimental to other specialized species, such as the hypogaeic species Hypoponera eduardi characteristic of the pre-eradication assemblages of this site (CA1).

While our results clearly showed the benefits of Carpobrotus eradication on ant occupancy at CA1 located on the cliff top and surrounded by low matorral, there was less evidence of benefits at CA2. This coastal site had patchy vegetation with some native species before eradication and higher rock cover, which remained at the same level throughout the study (Buisson et al. 2021). The eradication of Carpobrotus spp. probably did not trigger enough changes to the ecosystem to induce modifications in ant assemblage composition and structure.

Carpobrotus spp. eradication, and thus habitat change, was a stronger driver of change in the ant assemblages than climate. Our results suggest that the restoration of diversified ant assemblages, with higher foraging activity, was more effective at CA1, where Carpobrotus was denser and thicker. Increased ant activity goes hand-in-hand with strong habitat transformation and the return of native and diversified plant communities. Higher levels of multifunctionality are reached in species-rich ecosystems (Lefcheck et al. 2015) and many functions also depend on abundance (Soliveres et al. 2016). As ecosystem engineers, ants are known to participate in seed dispersal, and to improve soil structure and nutrient cycling (Folgarait 1998). Thus, increasing ant activity and richness at eradication sites may contribute to an increase in functions performed, including the enhancement of soil quality (Nkem et al. 2000) and seed-dropping frequency (Arnan et al. 2010). Ants play an active role in ecosystem recovery, and their quick response to the eradication here points to their efficacy as bioindicators when assessing restoration success.

Our data come from eradications carried out on a single small Mediterranean island, and this study should therefore be replicated on other islands to confirm our results. The harmful effects of Carpobrotus are leading to an increasing number of eradication campaigns, particularly in the Mediterranean. Ant monitoring would appear to be a relevant means of monitoring such restoration and conservation strategies.

Acknowledgements

We thank Port-Cros National Park and Naturalia-Environment for providing financial support for the study. We are grateful to all those who helped with field work and sorting insects from traps, as well as to Marjorie Sweetko for English language revision.

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Clarke KR (1993) Non-parametric multivariate analyses of changes in community structure. Australian Journal of Ecology 18:117–143. https://doi.org/10.1111/j.1442-9993.1993.tb00438.x
Courchamp F, Caut S, Bonnaud E, et al (2011) Eradication of alien invasive species: surprise effects and conservation successes. In: Island Invasives: Eradication and Management. pp 285–289
Cros S, Cerdá X, Retana J (2016) Spatial and temporal variations in the activity patterns of Mediterranean ant communities. Écoscience
Davidson DW, Samson DA, Inouye RS (1985) Granivory in the Chihuahuan Desert: Interactions within and between Trophic Levels. Ecology 66:486–502. https://doi.org/10.2307/1940397
Dawson W, Moser D, van Kleunen M, et al (2017) Global hotspots and correlates of alien species richness across taxonomic groups. Nat Ecol Evol 1:1–7. https://doi.org/10.1038/s41559-017-0186
De Almeida T, Blight O, Mesléard F, et al (2020) Harvester ants as ecological engineers for Mediterranean grassland restoration: Impacts on soil and vegetation. Biological Conservation 245:108547. https://doi.org/10.1016/j.biocon.2020.108547
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Cerdá X, Retana J, Manzaneda A (1998) The role of competition by dominants and temperature in the foraging of subordinate species in Mediterranean ant communities. Oecologia 117:404–412. https://doi.org/10.1007/s004420050674

Chenot J, Affre L, Gros R et al (2018) Eradication of invasive Carpobrotus sp.: effects on soil and vegetation. Restor Ecol 26:106–113. https://doi.org/10.1111/rec.12538

Clarke KR (1993) Non-parametric multivariate analyses of changes in community structure. Aust J Ecol 18:117–143. https://doi.org/10.1111/j.1442-9993.1993.tb00438.x

Courchamp F, Caut S, Bonnaud E et al (2011) Eradication of alien invasive species: surprise effects and conservation successes. In: Island Invasives: Eradication and Management. pp 285–289

Cros S, Cerdá X, Retana J (2016) Spatial and temporal variations in the activity patterns of Mediterranean ant communities. Écoscience

Davidson DW, Samson DA, Inouye RS (1985) Granivory in the Chihuahuan Desert: Interactions within and between Trophic Levels. Ecology 66:486–502. https://doi.org/10.2307/1940397

Dawson W, Moser D, van Kleunen M et al (2017) Global hotspots and correlates of alien species richness across taxonomic groups. Nat Ecol Evol 1:1–7. https://doi.org/10.1038/s41559-017-0186

De Almeida T, Blight O, Mesléard F et al (2020) Harvester ants as ecological engineers for Mediterranean grassland restoration: Impacts on soil and vegetation. Biol Conserv 245:108547. https://doi.org/10.1016/j.biocon.2020.108547

Deitch M, Sapundjieff M, Feirer S (2017) Characterizing Precipitation Variability and Trends in the World’s Mediterranean-Climate Areas. Water 9:259. https://doi.org/10.3390/w9040259

Del Toro I, Ribbons RR, Pelini SL (2012) The little things that run the world revisited: a review of ant-mediated ecosystem services and disservices (Hymenoptera: Formicidae). Myrmecol News 17:133–146

Fernández-Palacios JM, Kreft H, Irl SDH et al (2021) Scientists’ warning – The outstanding biodiversity of islands is in peril. Global Ecol Conserv 31:e01847. https://doi.org/10.1016/j.gecco.2021.e01847

Folgarait PJ (1998) Ant biodiversity and its relationship to ecosystem functioning: a review. Biodivers Conserv 7:1221–1224

Gordon DM (1991) Behavioral Flexibility and the Foraging Ecology of Seed-Eating Ants. Am Nat 138:379–411. https://doi.org/10.1086/285223

Gotelli NJ, Ellison AM, Dunn RR, Sanders NJ (2011) Counting ants (Hymenoptera: Formicidae): Biodiversity sampling and statistical analysis for myrmecologists. College of Arts and Sciences Faculty Publications

Herbers JM (1989) Community structure in north temperate ants: temporal and spatial variation. Oecologia 81:201–211. https://doi.org/10.1007/BF00379807

Hill JG, Summerville KS, Brown RL (2008) Habitat Associations of Ant Species (Hymenoptera: Formicidae) in a Heterogeneous Mississippi Landscape. Environ Entomol 37:453–463. https://doi.org/10.1093/ee/37.2.453

Hoffmann BD, Griffiths AD, Andersen AN (2000) Responses of ant communities to dry sulfur deposition from mining emissions in semi-arid tropical Australia, with implications for the use of functional groups. Austral Ecol 25:653–663. https://doi.org/10.1111/j.1442-9993.2000.tb00071.x

Howald G, Donlan CJ, Faulkner KR et al (2010) Eradication of black rats Rattus rattus from Anacapa Island. Oryx 44:30–40. https://doi.org/10.1017/S003060530999024X

Hulme PE, Brundu G, Camarda I et al (2008) Assessing the risks to Mediterranean islands ecosystems from alien plant introductions. 39–56

Jones HP, Holmes ND, Butchart SHM et al (2016) Invasive mammal eradication on islands results in substantial conservation gains. Proceedings of the National Academy of Sciences 113:4033–4038. https://doi.org/10.1073/pnas.1521179113

Kaspari M, Majer JD (2000) Chap. 7: Using Ants to Monitor Environmental Change. Ants - Standards Methods for Measuring and Monitoring Biodiversity. Smithsonian Institution Press, pp 89–98

Kier G, Kreft H, Lee TM et al (2009) A global assessment of endemism and species richness across island and mainland regions. Proceedings of the National Academy of Sciences 106:9322–9327. https://doi.org/10.1073/pnas.0810306106

Krebs E, Abba A, Gillet P et al (2015) Responses of reptile populations to the eradication of the Roof Rat (Rattus rattus) on Bagaud Island (Port-Cros National Park, Var, France). Rev Ecol-Terre Vie 70:99–109

Lasmar CJ, Bishop TR, Parr CL et al (2021) Geographical variation in ant foraging activity and resource use is driven by climate and net primary productivity. J Biogeogr 48:1448–1459. https://doi.org/10.1111/jbi.14089

Lassau SA, Hochuli DF (2004) Effects of habitat complexity on ant assemblages. Ecography 27:157–164. https://doi.org/10.1111/j.0906-7590.2004.03675.x

Lazzaro L, Tondini E, Lombardi L, Giunti M (2020) The eradication of Carpobrotus spp. in the sand-dune ecosystem at Sterpaia (Italy, Tuscany): indications from a successful experience. Biologia 75:199–208. https://doi.org/10.2478/s11756-019-00391-z

Lebas C, Galkowski C, Blatrix R, Wegnez P (2016) Fourmis d’Europe occidentale. Delachaux et Niestl&#233

Lefcheck JS, Byrnes JEK, Isbell F et al (2015) Biodiversity enhances ecosystem multifunctionality across trophic levels and habitats. Nat Commun 6:6936. https://doi.org/10.1038/ncomms7936

Lenda M, Witek M, Skórka P et al (2013) Invasive alien plants affect grassland ant communities, colony size and foraging behaviour. Biol Invasions 15:2403–2414. https://doi.org/10.1007/s10530-013-0461-8

Levings SC (1983) Seasonal, Annual, and Among-site Variation in the Ground Ant Community of a Deciduous Tropical Forest: Some Causes of Patchy Species Distributions. Ecol Monogr 53:435–455. https://doi.org/10.2307/1942647

Luque GM, Reyes López J (2007) Effect of experimental small-scale spatial heterogeneity on resource use of a Mediterranean ground-ant community. Acta Oecol 32:42–49. https://doi.org/10.1016/j.actao.2007.03.003

Majeed W, Khawaja M, Rana N (2021) Assessing fluctuation of ant populations in a distinct ecological habitat to track climate change effects. Biodiversitas J Biol Divers 22. https://doi.org/10.13057/biodiv/d220533

Majer JD (1983) Ants: Bio-indicators of minesite rehabilitation, land-use, and land conservation. Environ Manage 7:375–383. https://doi.org/10.1007/BF01866920

Martello F, de Bello F, de Morini MS C, et al (2018) Homogenization and impoverishment of taxonomic and functional diversity of ants in Eucalyptus plantations. Sci Rep 8:3266. https://doi.org/10.1038/s41598-018-20823-1

Martinez Arbizu P (2020) pairwiseAdonis: Pairwise multilevel comparison using adonis

Molinari N, D’Antonio C, Thomson G (2007) Carpobrotus as a case study of the complexities of species impacts. Ecosystem Engineers: Plants to Protists. Academic Press

Myers N, Mittermeier RA, Mittermeier CG et al (2000) Biodiversity hotspots for conservation priorities. Nature 403:853–858. https://doi.org/10.1038/35002501

Nascimento T, Oliveira N, Fagundes AI et al (2019) Diet selection of introduced black ratsRattus rattus L. in relation to plant availability on Berlenga Island, Portugal / Préférence alimentaire des rats noirs introduits Rattus rattus L. en relation avec la disponibilité de la flore sur l’île de Berlenga, Portugal. Ecologia Mediterranea 45:15–29. https://doi.org/10.3406/ecmed.2019.2072

Nkem JN, Lobry de Bruyn LA, Grant CD, Hulugalle NR (2000) The impact of ant bioturbation and foraging activities on surrounding soil properties. Pedobiologia 44:609–621. https://doi.org/10.1078/S0031-4056(04)70075-X

Novoa A, González L, Moravcová L, Pyšek P (2013a) Constraints to native plant species establishment in coastal dune communities invaded by Carpobrotus edulis: Implications for restoration. Biol Conserv 164:1–9. https://doi.org/10.1016/j.biocon.2013.04.008

Novoa A, Rodríguez R, Richardson D, González L (2013b) Soil quality: a key factor in understanding plant invasion? The case of Carpobrotus edulis (L). Biol Invasions 16:429–443. https://doi.org/10.1007/s10530-013-0531-y

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Table 1 : Site acronym with type and detailed description, main plant species identified by Buisson et al. (2021), and density of black rats (Rattus rattus) recorded by Ruffino et al. (2015)

Site name	Type of site	site Details	Vegetation	Rat density (individuals/ha) (Ruffino et al., 2015)
CA1	Carpobrotus eradication	Inland 0.5 ha patch of Carpobrotus aff. acinaciformis located on top of a 40 m high cliff	Low matorral vegetation surrounding the Carpobrotus patch. Recolonizing vegetation dominated by Lotus creticus subsp. cytisoides, Sonchus asper subsp. glaucescens, Bromus diandrus subsp. diandrus	between 46 and 60
CA2	Carpobrotus eradication	Coastal area with Carpobrotus edulis intermingled with small halophilous vegetation and rock patches	Coastal halophilous vegetation and low matorral inland. Main species are Lotus creticus subsp. cytisoides, Sonchus asper subsp. glaucescens, Frankenia sp.	below 37
RR1	Rat eradication	Nitrophilous grassland highly frequented by seagulls	Common species are Hordeum murinum subsp. leporinum, Sonchus asper subsp. glaucescens and Lotus creticus subsp. cytisoides	between 60 and 75
RR2	Rat eradication	Open low matorral vegetation	Small patches of grassland intermingled with shrubs Lotus creticus subsp. cytisoides, Pistacia lentiscus, and Frankenia sp.	between 46 and 60
CO1	Control	Low matorral vegetation	Shrubby vegetation with Pistacia lentiscus, Smilax aspera and Myrtus communis	below 37
CO2	Control	High matorral vegetation	Matorral vegetation with trees such as Phillyrea sp., Pistacia lentiscus and Erica arborea.	below 37

Table 2 : Summary of the main weather data from the Porquerolles weather station (9 kilometers from Bagaud island) during the four sampling periods (April to July and September to October)

	2010	2011	2013	2015	2017	2019	Mean over the six years
Mean maximum temperature (°C)	22.65	25.52	22.60	24.70	23.70	25.22	24.07
Mean minimum temperature (°C)	13.64	14.31	13.94	14.61	14.16	14.7	14.23
Annual cumulated precipitation (mm)	298.4	34.5	165.2	180.7	65.7	104.3	141.5

No competing interests reported.

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Mixed responses of ant assemblages to Carpobrotus spp. and Rattus rattus eradications on a small Mediterranean island

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Abstract

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1. Introduction

1. Materials and methods

1.1 Study area

1.2 Study sites

1.3 Sampling design

1.4 Data analysis

2. Results

2.1 Species richness and total abundance

2.2 Assemblage composition

2.3 Individual species abundance

3. Discussion

4. Conclusion

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Acknowledgements

References

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