Invasive Biology of Xenopus Laevis in Europe: Ecological Effects and Physiological Adaptations

MUSEUM NATIONAL D’HISTOIRE NATURELLE Ecole Doctorale Sciences de la Nature et de l’Homme – ED 227

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Spécialité : Ecologie

Présentée et soutenue publiquement par

Julien Courant Le 19 Septembre 2017

Invasive biology of Xenopus laevis in Europe: ecological effects and physiological adaptations

Sous la direction de : Monsieur Herrel, Anthony, Directeur de Recherche, Et Monsieur Secondi Jean, Maître de conférences JURY :

M. Clergeau, Philippe Professeur, Muséum national d’histoire naturelle, Paris (075) Président M. Herrel, Anthony Directeur de Recherche, Muséum national d’histoire naturelle, Paris (075) Directeur de Thèse M. Secondi, Jean Maître de conférences, Université de Lyon, Lyon (069) Directeur de Thèse M. Besnard, Aurélien Maître de conférences, CEFE-CNRS, Montpellier (034) Rapporteur M. Rödel, Mark-Oliver Professeur, Museum für Naturkunde, Berlin (Germany) Rapporteur M. Measey, John Chargé de Recherche, Stellenbosch University, Stellenbosch (South Africa) Examinateur M. Miaud, Claude Directeur de Recherche, Centre National de la Recherche Scientifique, Paris (075) Examinateur

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Table of contents

Résumé ...... 2

General Introduction ...... 6

Chapter I - Are invasive populations characterized by a broader diet than native populations?

...... 20 Introduction ...... 24 Material and Methods ...... 27 Results ...... 30 Discussion ...... 37

Chapter II - Assessing the impact of an invasive population of the African clawed frog on an amphibian community in western France.

...... 42 Introduction ...... 46 Material and Methods ...... 48 Results ...... 54 Discussion ...... 587

Second Part - Introduction ...... 64

Chapter III - Resources allocated to reproduction decrease at the range edge of an expanding population of an invasive amphibian.

...... 70 Introduction ...... 74 Material and Methods ...... 76 Results ...... 83

Discussion ...... 87

Chapter IV - Differences in mobility at the range edge of an expanding invasive population of Xenopus laevis in the west of France.

...... 92 Introduction ...... 96 Material and Methods ...... 98 Results ...... 105 Discussion ...... 110

Chapter V - Changes in dispersal allocation at the range edge of an expanding population.

...... 114 Introduction ...... 118 Material and Methods ...... 121 Results ...... 128 Discussion ...... 133

Chapter VI - Dynamic demography and time since colonization in an expanding population.

...... 138 Introduction ...... 142 Material and Methods ...... 146 Results ...... 151 Discussion ...... 156

General Discussion ...... 162

References ...... 176

Appendices ...... 210

Abstract ...... 307

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Figures list

Fig.Int.1 Photographical description of Xenopus laevis.

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Fig.Int.2 Global projection of the potential distribution of X. laevis.

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Fig.Int.3 General principle of the sampling design used during the studies performed on the invasive population of Xenopus laevis in western France.

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Fig.1.1 Two first principal components of the PCA representing the populations according to their diet.

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Fig.1.2 Occurrence of the main ecological traits among the five populations: terrestrial prey (black), benthic prey (dark blue), nektonic items (sky blue) and planktonic prey (cyan).

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Fig.1.3 Niche breadth calculated for diet data in native and colonized ranges of Xenopus laevis.

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Fig.1.4 Electivity index for each aquatic prey category consumed in South Africa (brown), Wales (dark-orange) and France (light-orange).

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Fig.2.1 Distribution of the expanding population of Xenopus laevis in western France, with the sites (in yellow) selected for the site-occupancy study.

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Fig.2.2 N-mixture estimates of species richness according to (a) the abundance of Xenopus laevis and (b) the distance to the introduction site. ix

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Fig.3.1 Location of the three pond clusters of Xenopus laevis sampled to analyze the variation in resource allocation to reproduction.

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Fig.3.2 Mean values (± Standard Error) of the SMIorgan of females Xenopus laevis captured in 2014.

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Fig.3.3 Reproductive investment of Xenopus laevis in the core, intermediate, and edge clusters as represented by the SMIorgan (mean ± SE) of females (left) and males (right).

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Fig.4.1 Current distribution of the invasive population of Xenopus laevis in the west of France.

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Fig.4.2 Top view of the track used to perform endurance tests.

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Fig.4.3 Snout–vent length for males and females from sites at the center and the periphery of the range.

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Fig. 4.4 Mean time until exhaustion for males and females from populations at the center versus the periphery of the range.

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Fig.5.1 Data sampling for the field survey (detailed in a and b) and the experiments.

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Fig.5.2 Experimental track built to estimate dispersal components.

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Fig.5.4 Estimations of dispersal distances for the range core and the range edge of the invasive populations of Xenopus laevis in Western France.

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Fig.6.1 Sampled areas for the Capture-Mark-Recapture survey and the skelettochronological study.

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Fig.6.2 Age structure estimates for the core (brown) and edge (orange) populations for females (filled symbols) and males (open symbols) according to the nonlinear regressions derived from the skeletochronological data.

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Fig.6.3 Demographic parameters, i.e. the survival probability (A) and the emigration probability (B), estimated using a multistate closed robust design method for males (open triangles) and females (filled triangles) from the core (brown) and the edge (orange) populations.

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Fig.D.1 Relative allocation of resources (filled circle for high, unfilled circle for low) to the life history traits at the range core (full line) and the range edge (dotted line) according to predictions by Burton et al., (2010) in blue and to our results in black.

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Tables list

Tab.1.1 Characteristics of the methods used to capture and describe the diet of Xenopus laevis.

...... 28 Tab.1.2 Relative abundance (in %) of the prey items identified in the native and colonized ranges of Xenopus laevis.

...... 32 Tab.2.1 Summary of the best models obtained with the N-mixture method (Royle, 2004) to estimate species richness.

...... 54 Tab.2.2 Effects of Xenopus laevis on the niche position and breadth of native amphibians and its co- occurrences with these native amphibians.

...... 57 Tab.3.1 Results of the linear mixed models performed to test the spatial variation in reproductive allocation across the colonized range of Xenopus laevis in France.

...... 84 Tab.3.2 Mean values and Standard Errors obtained for each sex for the SMI of gonads, adipose tissue and entire reproductive organ in the three clusters.

...... 85 Tab.4.1 Morphometric traits (means±s.e.) measured for each site and sex.

...... 100 Tab.4.2 Mean performance trait and body temperature at the end of the trial.

...... 103 Tab.4.3 Results of univariate analyses testing for differences in SVL and body temperature at the end of the stamina trial between populations and sexes.

...... 105

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Tab.4.4 MANCOVA performed on the morphometric data from sites 1 and 2 with SVL as covariate.

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Tab.4.5 MANCOVA performed on the morphometric data from sites 3 and 4 with SVL as covariate.

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Tab.4.6 MANCOVA performed on the performance traits with SVL and temperature as covariates.

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Tab.5.1 Model output for the linear mixed models run to test the effect of the cluster on REXPE and

DEXPE, taking into account environmental parameters.

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Tab.6.1 Parameters of the nonlinear regressions of the age of individuals by their snout-vent length.

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Tab.6.2 Estimates of population sizes and recapture rates according to the Capture-Mark-Recapture survey.

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Tab.6.3 Capture probabilities parameters estimated using the multistate open robust design method.

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Remerciements

Je tiens tout d‟abord à remercier les chercheurs qui ont accepté avec enthousiasme de siéger à mon jury de thèse : Aurélien Besnard et Mark-Oliver Rödel and tant que rapporteurs, et

Philipe Clergeau, Claude Miaud et John Measey en tant qu‟examinateurs.

Je remercie également chaleureusement mes directeurs de thèse, Anthony Herrel et Jean

Secondi, pour la confiance qu‟ils m‟ont accordée dans cette aventure, pour les nombreuses discussions que nous avons échangées. Je les remercie aussi pour m‟avoir laissé travailler avec une certaine liberté dans l‟organisation du travail de terrain et de laboratoire, condition sans laquelle mon épanouissement n‟aurait pas pu être aussi complet.

Les données utilisées pendant ma thèse provenant en majorité du travail de terrain, il est important de remercier les acteurs locaux de la conservation de la biodiversité pour leur aide, leur conseils ou tout simplement pour les échanges que nous avons eu au cours de ces trois années de terrain. Parmi ces acteurs, je tiens à citer la Communauté de communes du

Thouarsais en la personne de Rodolphe Olivier, l‟Agglomération du Bocage Bressuirais,

Sylvie Desgranges de la LPO Anjou, Bastien Richard du Parc Naturel Régional Loire-Anjou-

Touraine, Dorian Angot du Service Biodiversité de Chalonnes-sur-Loire et Myriam

Labadesse de la Société Herpétologique de France.

Il serait impossible de lister tous les propriétaires terriens que j‟ai croisés pendant ces trois années de terrain, mais je tiens tout de même à remercier la très grande majorité d‟entre eux qui m‟a accordé l‟accès à l‟une des quelques 300 mares visitées pendant la thèse. Je remercie tout particulièrement les membres des GAEC du Sault et du GAEC du Layon à pour m‟avoir laissé mener mes suivis de Capture-Marquage-Recapture sur leurs terrains pendant trois ans.

De nombreuses personnes sont venues m‟aider sur le terrain, je souhaiterais les remercier, dans un désordre total et une très probable incomplétude (toutes mes excuses aux oubliés), xvii pour leur temps et pour les bons moments passés ensemble malgré les conditions parfois précaires et la météo et/ou les vaches parfois hostiles : Damien, Célia, Vinciane, Matthieu,

Amandine, Vincent, Sophie, Jean-Marc, Julie, Sarah, Mélanie, Maman, Anthony, Jean,

Sandra, Hugo, Mélissa, Emilie.

Je tiens à remercier l‟ensemble des membres des deux équipes de mes laboratoires de l‟Université d‟Angers et du Muséum National d‟Histoire Naturelle de Paris, qui ont toujours

été d‟un contact facile et agréable. Le nombre de bons moments (à Angers au cours des raclettes et des « contests » où mauvaise foi et bonne humeur faisaient si bon ménage, ou des apéritifs et autres pots et événements à Paris) est incalculable et a permis de maintenir les liens entre les membres des équipes et tout particulièrement des doctorants. Je remercie à cet

égard Alexandra, Marion et Ameline pour l‟organisation des événements au labo de Paris et pour leur investissement permanent dans la vie de l‟équipe. Toujours au sein de ce labo, je tiens à remercier les trois gestionnaires, Nadine, Yamso et Manuela qui, malgré mon incapacité chronique à anticiper les échéances administratives, ont toujours été présentes et d‟une efficacité sans faille dans les différentes tâches.

Mon projet de thèse faisait partie d‟un projet de recherche européen, au sein duquel les différents membres ont pu échanger et travailler ensemble. Je remercie chacun des membres du projet INVAXEN : Charlotte De Busschere, Thierry Backeljau, Rui Rebelo, Raquel

Marques, Solveig Vogt, Dennis Rödder, Flora Ihlow, André De Villiers et John Measey. Je remercie également, Biodiversa, l‟institution européene qui a financé ma thèse et l‟intégralité du projet de recherche INVAXEN.

Les stagiaires avec qui j‟ai pu travailler directement pendant ma thèse ont tous fait preuve d‟une motivation, d‟une curiosité et d‟une capacité d‟investissement qui n‟ont cessé de m‟impressionner. Je vous en remercie, par ordre chronologique : Hugo Harbers, Viviane

Bereiziat, Elise Vollette, Lucille Guillemet, Vivien Louppe et Layla Adil. Je remercie

également les étudiants avec qui j‟ai pu travailler, ou simplement échanger, durant la thèse.

Il y a ceux qui ont permis que ma thèse se passe bien, et il y a ceux qui ont permis que je l‟envisage.

A cet égard, je me dois de remercier en premier lieu mes parents qui, m‟ont toujours laissé poursuivre mes études, qui m‟y ont aidé financièrement et logistiquement, et ce malgré le fait que personne (surtout pas moi) n‟était en mesure de leur dire où cela pourrait bien me mener.

Je n‟aurais jamais pensé pouvoir faire cette thèse avant ma rencontre avec Jean-Marc Thirion.

Cette rencontre a littéralement changé mes projets de vie personnelle et professionnelle, et je tiens à lui transmettre toute ma reconnaissance. Il est évident que je n‟en serai pas là aujourd‟hui si je n‟avais pas vécu ces stages à OBIOS et je ne suis pas près d‟oublier d‟où je viens.

Je remercie également Sarah, qui m‟a permis d‟acquérir une certaine confiance en mes capacités, qui m‟a encouragé à m‟engager dans cette thèse. Elle a surtout supporté ces moments insensés que les doctorants sont capables d‟imposer à leurs proches. Rien n‟effacera tout cela.

Mes derniers remerciements, mais non les moindres, vont à Juliette, qui a été présente pour moi malgré le contexte compliqué qui nous entourait. L‟achèvement de cette thèse aurait été une lugubre période sans sa compréhension, sa patience, son humour, et son amour.

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Résumé

Au niveau mondial, la compréhension des conséquences des menaces qui pèsent sur la biodiversité est un enjeu crucial pour la biologie de la conservation dans le contexte actuel de déclin de la biodiversité. Les espèces exotiques envahissantes comptent parmi ces grandes menaces mondiales, et peuvent localement induire de graves dommages sur les écosystèmes.

Etudier les phénomènes régissant les effets et l‟expansion potentielle de ces espèces apparait comme un élément primordial pour évaluer leurs impacts sur le long terme. Chaque espèce exotique envahissante induit en effet des impacts sur les écosystèmes qui lui sont propres, avec des particularités propres aux différentes populations envahissantes lorsque celles-ci sont multiples. Dans cette étude, nous focalisons nos efforts sur une population exotique envahissante du Xénope lisse, Xenopus laevis, en France, afin d‟apporter de nouvelles connaissances sur les interactions de cette population avec son environnement. Nous étudions

également, dans une seconde approche, les changements dans l‟allocation des ressources aux traits d‟histoire de vie, liés aux probabilités de reproduction, de survie et de dispersion durant l‟expansion de son aire de répartition. Cette seconde section vise à mieux connaître le potentiel de l‟espèce en termes d‟expansions futures.

Nous étudions le régime alimentaire de l‟espèce dans la population française et dans cinq autres populations exotiques envahissantes et naturelles. De cette étude, nous concluons que les populations pouvaient s‟étendre en consommant une diversité de proie pouvant être faible comme grande. Cette étude nous permet également de conclure que le régime alimentaire de

X. laevis se modifie d‟une population à l‟autre selon la disponibilité en proie, et que peu de taxons, hormis les catégories composées d‟animaux de petite taille (larve de diptères, zooplancton…) était positivement sélectionnés par cette espèce. Nous reportons ensuite, dans une étude de l‟occupation des amphibiens autochtones dans la population envahissante du

Xénope lisse en France, un impact significatif de l‟espèce sur la communauté native

d‟amphibiens. Les résultats de ces études tendent à renforcer les mentions publiées précédemment dans la littérature concernant l‟importance de l‟impact de cette espèce sur les

écosystèmes colonisés. Evaluer dans quelle mesure les populations exotiques envahissantes de cette espèce sont susceptibles de s‟étendre, et d‟induire ces effets néfastes sur les nouvelles zones colonisées a été l‟objet de la seconde section de cette thèse. Nous y reportons, par une comparaison de la taille des gonades entre les individus du centre de l‟aire de répartition et du front de colonisation, une diminution de l‟investissement dans la reproduction au front de colonisation. L‟investissement dans la dispersion montre un patron opposé, avec une augmentation de l‟allocation à ce trait sur le front de colonisation de la population en expansion. Nous étudions finalement la dynamique des populations et reportons une diminution de la survie et de la densité au centre de l‟aire de distribution. Ces résultats ont un grand intérêt pour la gestion des conséquences de ces expansions sur le long terme chez X. laevis. Cependant, ils présentent également un intérêt plus fondamental ; celui de ne pas suivre exactement les prédictions fournies par les études théoriques portant sur l‟évolution des compromis entre les traits d‟histoire de vie durant les expansions d‟aire de distribution.

Prendre en compte, dans ces études théoriques, des patrons d‟allocation des ressources qui diffèrent des patrons classiques, semble donc nécessaire en biologie de l‟évolution et de la conservation. La combinaison des résultats obtenus sur l‟impact actuel de l‟espèce sur son milieu en France et sur les changements d‟allocation des ressources suggère que le potentiel d‟impact au long terme de la population de X. laevis est important en France, mais aussi dans les autres zones où l‟espèce a été, ou est susceptible d‟être, introduite.

GENERAL INTRODUCTION

General Introduction

We are currently facing a strong decline of biodiversity, inducing rarefactions and extinctions of species in the wild (Butchart et al., 2010). This phenomenon is related to the increasing intensity of global human activity, resulting in increased chemical pollution (e.g. Barker &

Tingley, 1992; Maiti & Chowdhury, 2013; Munir et al., 2016; McMullin et al., 2016), global climate change (e.g. Thomas et al., 2004; Bálint et al., 2011; Bellard et al., 2012), over- exploitation, habitat loss and fragmentation (Pimm & Raven, 2000; Fahrig, 2003; Krauss et al., 2010; Essl et al., 2015), and the introduction of invasive species (e.g. Clavero et al., 2009;

Katsanevakis et al., 2014; Galiana et al., 2014; Early et al., 2016). Locally, biodiversity decline is often due to, and enhanced, by more than one threat, thus increasing the difficulty of establishing effective conservation actions (Brook et al., 2007).

Invasive species have been claimed as the second most important cause of the current decline in biodiversity in the USA, after over-exploitation and habitat loss (Wilcove et al., 1998). The relative importance of invasive species on the biodiversity decline at the global scale has since been questioned (Gurevitch & Padilla, 2004; Davis, 2011), and over-exploitation and habitat loss are currently accepted as being the principal threats (Millenium Ecosystem Assessment,

2005; WWF, 2014). However, studies on the relations between biodiversity decline and invasive species suggests that this threat can locally have harmful impacts on biodiversity

(e.g. Clavero et al., 2009; Katsanevakis et al., 2014; Galiana et al., 2014; Doherty et al.,

2016). To improve our understanding of the contribution of invasive species on the biodiversity decline at a global scale, more empirical studies on the ecological and economic impacts of these species are needed (Olson, 2006; Molnar et al., 2008; Pyšek et al., 2012;

Early et al., 2016).

Because the impact of an invasive species is partially related to its potential for expansion, understanding the ecological and evolutionary processes driving the expansion of its populations is of crucial importance to predict their long term effects (Sakai et al., 2001). General Introduction

These introductions moreover represent opportunities to study the evolutionary processes determining the success of expanding populations (Sax et al., 2007) and to identify the intrinsic mechanisms and extrinsic conditions allowing expansions (Phillips et al., 2008).

From a general viewpoint, populations shifting their ranges because of climate change or human induced introductions are of great interest for evolutionary and conservation biology

(Thomas et al., 2001; Chuang & Peterson, 2016). The success and extent of expanding populations are intrinsically driven by population parameters, i.e. growth rate, dispersal and density dependence. These parameters summarize the trade-offs between the life-history traits related to dispersal, reproduction, and competitive ability, driving the probability for each individual to disperse, reproduce, and survive (Burton et al., 2010), respectively.

According to theoretical studies (Phillips et al., 2010; Burton et al., 2010), resource allocation to dispersal and reproduction should increase at the range edge of an expanding population at the detriment of the resource allocation to traits related to competitive ability. This prediction is partially explained by the higher probability for long-distance dispersers at the range edge to reach a breeding site where access to resources and intra-specific competition is reduced

(Brown et al., 2013). Higher dispersal has been documented at the range edge of many expanding populations, through a modified morphology (e.g. Haag et al., 2005; Phillips et al.,

2006; Hassal et al., 2009; Laparie et al., 2013), an increased performance (Hartman et al.,

1997; Phillips et al., 2006; Brown et al., 2007), or a change in exploratory behaviour

(Llewellyn et al., 2010; Lliebl & Martin, 2012; Myles-Gonzalez et al., 2015).

At the range core, there is a higher probability of occupying a breeding site where the conspecific density is high (Baguette & Van Dyck, 2007), resulting in the selection of individuals with higher competitive ability (Burton et al., 2010) at the range core compared to the range edge. This prediction has been corroborated by empirical results (e.g. Llewellyn et al., 2012; Kolbe et al., 2014; Therry et al., 2014). Besides the changes in the allocation to

9 other traits like dispersal and competitive ability, the resource allocation to reproduction is also influenced by the Allee effect. This effect is defined as the decline in the fitness of individuals when population size and density is low, possibly resulting in population decline or extinction (Courchamp et al., 2008). Allocation to reproduction is enhanced at the range edge when the population does not locally suffer from the consequence of the Allee effect, and it is reduced when the population faces this phenomenon (Burton et al., 2010; Chuang &

Peterson, 2016). Empirical studies focusing on reproductive allocation have demonstrated that both enhanced (Hill et al., 1999; Rebrina et al., 2015) and reduced (Hughes et al., 2003;

Amundsen et al., 2012; Colautti & Barret, 2013) allocation to reproduction occurs at the range edge of expanding populations. Several studies have confirmed a relation between the Allee effect and the reduced reproductive success at the range edge of expanding populations

(Contarini et al., 2009; Gutowsky & Fox, 2011; Miller & Inouye, 2013).

These changes in trade-offs are likely driven by evolutionary mechanisms such as natural selection, or – more likely – spatial sorting (Shine et al., 2011). This phenomenon is described in the context of range expansions, and can be defined as the accumulation, at the expanding range edge, by repeated mating between fast-dispersing individuals, of traits enhancing the dispersal rate of the population, regardless how this phenotype can affect the survival and reproductive success of the organisms (Shine et al., 2011). It has first been suggested in a study on the cane toad, Rhinella marina, an anuran invading wide areas in Northern Australia

(Phillips et al., 2008). Further studies remain necessary to assess the actual role of spatial sorting in evolutionary processes during range expansions beyond the case of R. marina.

Local adaptation is likely favoured if a population that expands into different landscapes and/or in areas submitted to different climate conditions. Thus, selecting an expanding population occupying a homogenous area submitted to only one climatic context would seem General Introduction judicious to assess the effect of spatial sorting on changes in trade-offs between life-history traits.

During range expansions, the allocation of resources to life-history traits is also dependent on extrinsic factors, such as the presence of strong competitors and resource availability (Burton et al., 2010). Thus, to provide accurate predictions about the impact of an invasive population on the long term, both ecological knowledge about the interactions of this population with the colonized environment and the evolutionary phenomena driving its expansion are necessary.

In this work, I focused my efforts on studying an invasive and expanding population to provide the necessary knowledge to assess its potential long term impacts. I first analyzed the interactions of this population with the environment it colonizes by studying its dietary requirements and its effects on the native communities. I then focused on the resource allocation changes occurring at the range edge of this population, and at the mechanisms that might drive these changes. Selecting a model for this study necessitated to choose an expanding population occupying a homogeneous area in terms of climate and landscape. The population of the African clawed frog (Fig.Int.1), Xenopus laevis (Daudin, 1802), introduced in France, appeared as an interesting model for this study.

This species has been used since the first half of the twentieth century, for human pregnancy testing (Shapiro & Zwarenstein, 1934) and as a biological model in developmental biology

(Gurdon & Hopwood, 2000; van Sittert & Measey, 2016). As a consequence of the trade and farming providing research laboratories in addition with the interest of the aquarium trade for this species (Weldon et al., 2007), X. laevis was introduced in several localities on four continents (Measey et al., 2012). In France, the African clawed frog has been introduced during the 1980s, after escape events from a breeding facility (Fouquet, 2001). It has then

11 expanded its range to reach the Layon and the Thouet, two rivers located in the Loire valley, and finally reached the Loire river.

Fig.Int.1 Photographical description of Xenopus laevis. a) Picture of a male captured in France (scale 1:1); b) illustration of the sexual dimorphism with the female on the left, on average a third longer than a male; c) sexual characteristics allowing the identification of each sexe. Females, during their period of sexuality, exhibit a rounded cloaca, while males exhibit a flat one. Males also carry dark nuptial callosities on their forelimb during the breeding period.

It has been classified as the amphibian with the second strongest impact on native species in the world, with the cane toad, R. marina, as the first species on this list (Measey et al., 2016).

Even if X. laevis has already been suggested to negatively impact native amphibians in an invaded area (Lillo et al., 2011), more detailed studies remain necessary to understand the ecological effects of this invasive species on the colonized environment, and not only on amphibians (Measey et al., 2016; Kumschick et al., 2017).

The African clawed frog has long been considered as a fully aquatic anuran (Casterlin &

Reynolds, 1980). This may be partially due to the possibility to keep it in captivity in fully General Introduction aquatic conditions (Shapiro & Zwarenstein, 1934). Its hind limb, very well adapted to an aquatic life style (Videler & Jorna, 1985; Richards, 2010), and its diet, essentially composed of aquatic invertebrates (e.g. McCoid & Fritts, 1980; Measey, 1998; Amaral & Rebelo, 2012;

Courant et al., 2014 [Appendix A]), suggest that X. laevis is indeed fully aquatic. But it is also known to use overland movements, even if empirical data about this phenomenon are still rare

(Measey, 2016). Moreover, the potential for expansion in this species is large (Fig.Int.2, especially in Europe, and under various scenarios of climate change (Ihlow et al., 2016

[Appendix B]).

In France, eradication programs have been put in place to limit the expansion of the African clawed frog. These programs mainly concern the southern part of the population (in the Deux-

Sèvres department), where institutions reacted quickly to the discovery of the species in the area and the recommendations of local naturalist associations (Fouquet, 2001; Grosselet et al.,

2005; Thirion et al., 2009). Until recently, far less attention has been devoted to the northern axes of the population range (in the Maine-et-Loire department), leading to a less accurate historical knowledge of its colonization by X. laevis.

Fig.Int.2 Global projection of the potential distribution of X. laevis A) derived from the maximum entropy SDM; B) derived from the ensemble SDM. Probability ranging from moderate (dark blue) to highly suitable (yellow). Source: Ihlow et al., 2016 [Appendix B].

General Introduction

Despite the occurrence of invasive populations of X. laevis on four continents and the growing number of studies addressing this issue, knowledge about its ecology and the evolutionary processes driving the expansions is still poor and deserves more attention in the context of conservation and eco-evolutionary biology. Beyond the case of X. laevis, understanding the changes in resource allocation in such a recently expanding population may shed light on the mechanisms ruling this phenomenon and driving these trade-offs.

To address these issues, I used a common sampling principle (Fig.Int.3). Ponds and individuals were compared according to the location of the ponds in the range in western

France. Core and edge individuals were compared, using as many spatial replicates as it was possible for each study. I also used ponds in intermediate areas when the study necessitated, thus providing a more accurate measure of the trait or phenomenon that was analysed. When necessary, i.e. to study the interactions of X. laevis with its environment, ponds were also sampled out of the colonized range.

Fig.Int.3 General principle of the sampling design used during the studies performed on the invasive population of Xenopus laevis in western France.

General Introduction

The following chapters are organized in two main parts. In the first part, dedicated to the ecological interactions of X. laevis with its environment, I first document the dietary diversity and requirements of this species in native and in several invasive populations. In the second chapter, I propose a study of the effects of X. laevis on the community of amphibians in the colonized area in western France. In the second part of this work, I study the evolutionary processes driving the expansion of X. laevis in France. Thus, the third chapter focuses on the allocation of resources to reproduction while the fourth one consists of a comparison of locomotor capacities of individuals between the range core and the range edge of the population introduced in France. In the fifth chapter, I perform a thorough analysis of dispersal allocation by studying the movements of free-ranging individuals based on a field survey and an experimental approach. In the last chapter, I perform a study on the comparison of population dynamics between the range core and the range edge of the invasive X. laevis population in France using a capture-mark-recapture survey. Finally, I synthetically discuss these results, highlight the insights they bring, and the relevance of their use for extrapolation beyond the case of the X. laevis population I studied and beyond the case of X. laevis itself.

Chapter I

CHAPTER I

Are invasive populations characterized by a broader diet than native populations?

Chapter I

Chapitre I. Les populations exotiques envahissantes sont-elles caractérisées par un régime alimentaire plus varié que les populations natives ?

Courant J, Vogt S, Marques R, Measey J, … Herrel A (2017) Are invasive populations characterized by a broader diet than native populations? PeerJ 5:e3250; DOI 10.7717/peerj.3250 [Appendix C]

Résumé

Les populations exotiques envahissantes comptent parmi les plus importantes menaces pour la biodiversité. Les espèces exotiques envahissantes sont souvent caractérisées par un régime alimentaire diversifié, leur permettant de prospérer dans une gamme d‟environnements variés. Ainsi, comparer le régime alimentaire de populations exotiques envahissantes avec celui des populations autochtones d‟une même espèce est un élément important pour comprendre ses besoins, sa flexibilité en termes de diversité de ressources, et ses impacts associés sur l‟environnement colonisé. Dans la présente étude, nous avons comparé le régime alimentaire de populations du Xénope lisse, Xenopus laevis, dans son aire naturelle, avec celui de populations introduites et envahissantes. Chaque catégorie de proie détectée dans les contenus stomacaux a été assignée à une catégorie écologique (benthique, pélagique, planctonique, terrestre), permettant ainsi une comparaison de la diversité des traits écologiques parmi les proies dans le régime alimentaire des populations étudiés. La comparaison des régimes alimentaires a aussi été réalisée par l‟estimation d‟un indice de taille de niche trophique pour chaque population, et d‟un indice de sélectivité des proies pour trois des six populations étudiées. Nos résultats ont montrés que la taille de niche pouvait aussi bien être étroite que large dans les populations exotiques envahissantes. D‟après le calcul de sélectivité, le zooplancton était favorisé dans la plupart des populations. Les proportions de proies des différents traits écologiques et la variabilité du régime alimentaire au sein des différentes populations, et entre elles, suggèrent que X. laevis est un prédateur généraliste dans les populations naturelles comme dans les populations exotiques envahissantes. Des déplacements de niches trophiques ont été observes dans cette étude, et semblent liés à la disponibilité en ressources. Le Xénope lisse est susceptible d‟induire un impact sur les écosystèmes aquatiques, à cause de son occupation quasi permanente des milieux aquatiques et de son importante consommation de taxons clés pour les réseaux trophiques des milieux aquatiques tels que les mares. Chapter I

Chapter I. Are invasive populations characterized by a broader diet than native populations?

Courant J, Vogt S, Marques R, Measey J, … Herrel A (2017) Are invasive populations characterized by a broader diet than native populations? PeerJ 5:e3250; DOI 10.7717/peerj.3250 [Appendix C]

Abstract

Invasive species are among the most significant threats to biodiversity. The diet of invasive animal populations is a crucial factor that must be considered in the context of biological invasions. A broad dietary spectrum is a frequently cited characteristic of invasive species, allowing them to thrive in a wide range of environments. Therefore, empirical studies comparing diet in invasive and native populations are necessary to understand dietary requirements, dietary flexibility, and the associated impacts of invasive species. In this study, we compared the diet of populations of the African clawed frog, Xenopus laevis in its native range, with several areas where it has become invasive. Each prey category detected in stomach contents was assigned to an ecological category, allowing a comparison of the diversity of ecological traits among the prey items in the diet of native and introduced populations. The comparison of diets was also performed using evenness as a niche breadth index on all sampled populations, and electivity as a prey selection index for three out of the six sampled populations. Our results showed that diet breadth could be either narrow or broad in invasive populations. According to diet and prey availability, zooplankton was strongly preferred in most cases. In lotic environments, zooplankton was replaced by benthic preys, such as ephemeropteran larvae. The relative proportions of prey with different ecological traits, and dietary variability within and between areas of occurrence, suggest that X. laevis is a generalist predator in both native and invasive populations. Shifts in the realized trophic niche are observed, and appear related to resource availability. Xenopus laevis may strongly impact aquatic ecosystems because of its near complete aquatic lifestyle and its significant consumption of key taxa for the trophic relationships in ponds.

Introduction

Invasive species usually occupy a wide geographical range in their native area. Invasive species are typically characterized by a number of traits that favor the establishment and spread across new ecosystems, including a broad environmental tolerances, high genetic variability, rapid growth, early sexual maturity combined with a high reproductive rate, short generation time, broad diet, gregariousness, rapid dispersal, and they are often commensal

(Ricciardi & Rasmussen, 1998). Of course not all invasive species meet all these criteria

(Lodge, 1993). For example, successful invaders do not necessarily exhibit a broad diet

(Vazquez, 2006). Yet, a large dietary niche breadth is frequently considered as a hallmark of an invasive taxon.

The dietary niche is a component of the Eltonian niche, defined as the position of an organism, exhibiting a null population growth rate, in the trophic relationships with others organisms of the ecosystem such as its nutrients, predators and competitors (Chase & Leibold,

2003). Another aspect of the ecological niche is the Grinellian niche, defined as the set of all values of the abiotic parameters enabling the occupancy of an area by a species (Soberón,

2007). The Grinellian niche has known a recent intensification of research with the development of species distribution models (e.g. Angetter et al., 2011; Guisan et al., 2013).

Based on a set of occurrence records and predictor variables, these models determine a given species fundamental niche and facilitate the assessment of potential niche shifts when projected onto novel conditions. While these models are based on the assumption that species retain their ancestral traits over time (see Ackerly, 2003; Wiens & Graham, 2005) recent evidence (Broennimann et al., 2007; Mukherjee et al., 2012; Stiels et al., 2015) revealing shifts in realized Grinellian niches on a macro-ecological scale call this concept into question.

How far Eltonian niches operating on a population level are variable, under the assumption of niche conservatism, is less well studied. Whether a population maintains its characteristics, or Chapter I shifts them in the course of the invasion process will contribute determining its ecological impact. This information is therefore of crucial importance for conservation practitioners facing the threat posed by invasive species.

Native to southern Africa, the African clawed frog, Xenopus laevis, has been introduced in many countries on four continents, where accidentally and deliberately released individuals have established viable populations (see Measey et al., 2012). Despite its importance as model biological organism (Cannatella & De Sa, 1993), and the abundance of invasive populations, few field studies have been undertaken in colonized ranges (for a review, see

Measey et al., 2012). Xenopus laevis has been reported to negatively affect the invaded ecosystems, and as a consequence has been ranked as having the second greatest impact on native ecosystems by any amphibian (Measey et al., 2016). Shifts in the Grinellian niche of X. laevis have been recently demonstrated (Rödder et al., 2017), whereas studies on changes in the Eltonian niches have not yet been undertaken.

The diet of X. laevis has been studied in the species‟ native range of South-Africa (Schoonbee et al., 1992), as well as in several introduced populations in the United States of America

(USA) (McCoid & Fritts, 1980), Wales (Measey, 1998a), Chile (Lobos & Measey, 2002),

Italy (Faraone et al., 2008), Portugal (Amaral & Rebelo, 2012), and France (Courant et al.,

2014). In most studies, the majority of prey items are aquatic, with zooplankton and dipteran larvae being the most frequent. Measey (1998b) also noticed the importance of terrestrial prey. While stomach content analyses conducted in Portugal (Amaral & Rebelo, 2012) and in the USA (McCoid & Fritts, 1980) revealed that X. laevis consumed eggs of fishes and amphibians, no study has reported a direct impact linked to predation. A similar study, conducted in South African aquaculture ponds, revealed that farmed fish larvae constituted a large proportion (5 to 25 % occurrence frequency according to fish size) of frog stomach

25 contents (Schramm, 1987), while another study found X. laevis to consume large quantities of anuran eggs and larvae (Vogt et al., 2017).

Given the wide diversity of prey items, dietary studies on X. laevis usually suggest a generalist feeding behaviour, but only one study has thus far investigated prey electivity

(Measey, 1998a). No previous studies have explicitly compared the feeding behavior of populations in different ecological contexts, and with different invasion histories. In this study, we compiled published data on the diet of X. laevis from the USA (McCoid & Fritts,

1980), Wales (Measey, 1998a), Chile (Lobos & Measey, 2002), and South Africa (Vogt et al.,

2017) with data collected during recent field work in Portugal and France to test the hypotheses that (i) trophic niche breadth is wider in native populations, hence releaving the capacity of the species to readily adapt to novel environments; and (ii) the diet of invasive populations differs significantly between the invaded ranges depending on local prey availability, and thus resulting in a low degree of electivity and population-specific niche shifts.

Chapter I

Material & Methods

Data sampling

Our dataset comprised 1458 individuals from six countries, across four continents (Tab.1.1).

In most areas (Chile, South Africa, Wales, France), frogs were caught using funnel traps. In

Portugal, animals were captured using electrofishing because the colonized habitats, mainly fast flowing streams, prevented the use of traps. In South Africa, research permission was issued by CapeNature (AAA007-01867) and SANParks, with ethics clearance from

Stellenbosch University Research Ethics Committee: Animal Care & Use (SU-ACUD15-

00011). Animals from the Portuguese invasive population were captured under permit nº

570/2014/CAPT from Instituto da Conservação da Natureza e das Florestas, in the scope of the “Plano de erradicação de Xenopus laevis nas ribeiras do Concelho de Oeiras”

(Erradication program of the council of Oeiras for X. Laevis). In France, research permit was provided by the prefecture of the Deux-Sèvres department.

Stomach content samples were obtained either by stomach flushing or dissection, following euthanasia of individuals by lethal injection of sodium pentobarbital or immersion in MS222.

We considered that analyzing and comparing data collected with both dissection and flushing methods were valid and did not induce any bias (Wu et al., 2007).

Tab.1.1 Characteristics of the methods used to capture and describe the diet of Xenopus laevis.

*The population introduced in Wales went extinct twenty years after the data collection used in our

study (Tinsley et al., 2015). ** Geographical coordinates (WGS 84), northwestern (NW) and

southeastern corners (SE), of the minimum rectangle encompassing all sampled sites for NS>1.

South Africa Wales France Chile Portugal USA

Population status Native Extinct* Invasive Invasive Invasive Invasive

Period of capture From 06/2014 From 05/1995 From 05/2014 01/1998 From 06/2014 1975-1976

To-09/2014 To 08/1996 To 10/2014 03/2001 To 08/2014

Geographical coordinates Latitude/Longitude S 34°18‟24 „‟ N 51°27‟33‟‟ N 47°16‟14„‟ S 33°29‟ N 38°45‟09„‟ See McCoid

NW** E 18°25‟35‟‟ W 3°33‟11‟‟ W 0°33‟56‟‟ W 70°54‟ W 9°17‟27‟‟ & Fritts, 1980

Latitude/Longitude S 34°20‟06 „‟ N 46°53‟41„‟ S 33°37‟ N 38°42‟35„‟ See McCoid NA SE** E 19°04‟29‟‟ W 0°31‟11‟‟ W 70°39‟ W 9°16‟25‟‟ & Fritts, 1980

Sampling design Method Trap Trap Trap Trap Electrofishing NA

From 1 to 1 Capture occasion/site 29 3 1 From 1 to 4 4

Number of sites 8 1 26 2 12 1

Number of individuals 164 375 438 48 352 81

Prey Availability Yes Yes Yes No No No

Habitat Type Ponds Pond Ponds Ponds Streams Streams

Flushing/ Prey collection method Flushing Dissection Dissection Dissection Dissection Dissection

Published data Prey frequency in stomachs Yes Yes No Yes No Yes

Niche breadth Yes No No No No No

Electivity Yes Yes No No No No

Individual data Yes No Yes Yes Yes No Chapter I

Data analysis

Prey items retrieved from the stomach content samples were identified to the lowest taxonomic level possible. However, for analytical consistency, we retained the lowest common taxonomic level that could be identified for all prey items. As volume and mass of prey items were not available for most studies, analyses were performed using prey frequencies. Each taxonomic prey category was assigned to one of the following ecological traits: plankton, benthos, nekton and terrestrial. Some groups of invertebrates belong to different ecological trait categories depending on their life stage, e.g. aquatic in their larval stage and terrestrial in their adult stage (Diptera, Ephemeroptera, Trichoptera). Thus, adults and larvae were treated separately when assigned to different ecological traits even though they belong to the same taxonomic prey category.

The diet of populations was first compared by calculating the relative abundance of each prey category and ecological trait. The relative abundance of prey classes (aquatic invertebrates, terrestrial invertebrates, vertebrates) was also calculated. To assess variation in diet between populations, a Principal Component Analysis (PCA) was performed on reduced and centered relative abundances of each prey category.

The concept of niche breadth can be applied to comparative studies of diet between populations or species (Slatyer et al., 2013), even if it has not always been treated within this contextual vocabulary (e.g. Rehage et al., 2005; Luiselli et al., 2007; Dalpadado & Mowbray,

2013). We calculated niche breadth for all populations using the evenness measure J’. This index is based on the Shannon-Wiener‟s index H‟ (Shannon & Weaver, 1964), as recommended by Colwell & Futuyma (1971):

− ∑(p ∗ log p ) 퐽′ = i i log 푛

Where pi is the proportion of the prey category i in the diet and n is the number of food categories.

To test whether J’ is affected by the number of study sites the relationship of both variables was assessed using a nonlinear regression.

Prey availability was quantified in habitats for three of the six populations (France, Wales and

South Africa) included in this study. Following Measey (1998a), the same sampling method was applied to all populations. Prey selection was assessed using the Vanderploeg & Scavia‟s

(1979) relativized electivity index recommended by Lechowicz (1982):

1 푟푖 푊푖 − ∗ 푛 푝푖 퐸 = with 푊푖 = 1 ∑ 푟푖/푝푖 푊푖 + 푛

Where ri is the relative abundance of prey category i in the diet and pi is the relative abundance of prey category i in the environment. The number of prey categories included in the analysis is represented by n.

Results

Across all samples zooplankton was the most common prey type with a mean relative abundance of 56.21% (Standard Deviation = 32.80%), followed by ephemeropteran larvae

(10.31% ± 23.40%), dipteran larvae (9.68% ± 7.23%), and gastropods (7.24% ± 6.07%). The fifth and sixth most represented prey items were amphibian eggs, excluding X. laevis (4.86%

± 11.08%) and X. laevis eggs (3.46% ± 7.55%) respectively. Three aquatic invertebrate orders

(Coleoptera, Diptera, and Hemiptera) were detected in all study sites (Tab.1.2), while most of the terrestrial categories were exclusively found in one or two sites. Cannibalism of larvae and/or eggs was recorded in every locality, except Chile. Chapter I

Fig.1.1 Two first principal components of the PCA representing the populations according to their diet.

Tab.1.2 Relative abundance (in %) of the prey items identified in the native and colonized

ranges of Xenopus laevis. When prey items were observed in very low quantities (N<3), they were

noted as <0.01 % in the table. Below the name of each ecological category, we mentioned the mean

and the standard deviation of the relative abundance of items found in stomach contents.

Cat South

N° Africa Wales France Chile Portugal USA Aquatic invertebrates 65.99 99.15 80.68 98.54 96.56 98.13 1 Zooplankton 51.87 92.55 39.62 82.51 2.28 68.89 Benthos 8.86 6.45 37.08 14.85 93.90 27.41 34.80% ± 34.17% 2 Annelids 0.03 0.00 0.02 0.00 0.05 0.00 3 Turbellaria 0.00 <0.01 0.00 0.00 0.00 0.00 4 Gastropoda 0.00 0.00 9.82 10.64 15.05 7.99 5 Bivalvia 0.00 0.24 0.44 0.00 0.05 0.00 6 Acari 4.63 0.00 0.00 0.00 0.05 0.00 8 Amphipoda 1.83 0.02 3.09 0.00 0.05 4.21 9 Isopoda 0.00 0.00 0.00 0.00 2.62 0.05 10 Decapoda 0.0 0.00 0.00 0.00 0.06 0.00 11 Diptera (larvae) 1.78 6.14 20.90 4.21 15.24 9.91 12 Ephemeroptera (larvae) 0.00 0.05 2.79 0.00 58.02 0.98 13 Trichoptera (larvae) 0.59 0.00 0.02 0.00 2.71 4.27 Nekton 5.26 0.14 4.00 1.19 0.50 1.83 7.74% ± 13.87% 14 Coleoptera (larvae) 1.97 0.03 0.02 0.00 0.06 0.27 15 Coleoptera (adult) 0.67 0.09 1.74 0.44 0.38 0.34 16 Heteroptera 1.09 0.02 1.12 0.18 0.01 0.16 17 Zygoptera (larvae) 0.79 0.00 0.38 0.44 0.00 0.53 18 Anisoptera (larvae) 0.74 0.00 0.74 0.13 0.05 0.53 Terrestrial invertebrates 0.86 0.43 0.23 1.46 0.70 0.24 0.64% ± 0.46% 19 Arachnida 0.46 0.01 0.00 1.21 0.05 0.11 20 Isopoda 0.00 0.05 0.00 0.03 0.04 0.00 21 Chilopoda 0.00 0.00 0.00 0.00 0.01 0.00 22 Diplopoda 0.00 0.00 0.00 0.00 0.01 0.00 23 Diptera 0.09 0.06 0.00 0.00 0.04 0.08 24 Neuroptera 0.04 0.00 0.00 0.00 0.00 0.00 25 Hymenoptera 0.27 0.00 0.00 0.05 0.16 0.00 26 Coleoptera 0.00 0.01 0.05 0.00 0.00 0.05 27 Lepidoptera (larvae) 0.00 <0.01 0.00 0.00 0.00 0.00 28 Lepidoptera (adult) 0.00 0.00 0.00 0.15 0.00 0.00 29 Dermaptera 0.00 0.01 0.00 0.03 0.00 0.00 30 Heteroptera 0.00 0.00 0.02 0.00 0.10 0.00 31 Annelids 0.00 0.02 0.09 0.00 0.02 0.00 32 Orthoptera 0.00 0.00 0.08 0.00 0.00 0.00 33 Aphids 0.00 0.27 0.00 0.00 0.00 0.00 34 Trichoptera 0.00 0.00 0.00 0.00 0.02 0.00 35 Ephemeroptera 0.00 0.00 0.00 0.00 0.24 0.00 Vertebrates 33.13 0.41 19.09 0.00 2.74 1.63 9.45% ± 13.54% 36 Fish (adult) 0.00 0.00 0.11 0.00 0.02 0.08 37 Fish (egg) 0.00 0.00 0.00 0.00 0.82 0.00 38 Amphibia (adult) 0.14 0.00 0.02 0.00 0.00 0.00 3 Amphibia (larvae) 3.37 0.00 0.00 0.00 0.00 0.00 39 Amphibia (egg) 27.62 0.00 0.08 0.00 1.65 0.00 40 X. laevis (larvae) 0.04 0.00 0.08 0.00 0.01 0.03 41 X. laevis (egg) 0.00 0.41 18.81 0.00 0.00 1.52 42 Amphibia (rest) 1.98 0.01 0.00 0.00 0.22 0.00 43 Bird (feather) 0.00 <0.01 0.00 0.00 <0.01 0.00 44 Mammals 0.00 <0.01 0.00 0.00 0.00 0.00 Chapter I

Aquatic invertebrates represented the most consumed prey item class, with a relative abundance ranging from 66% in South Africa to 99% in Wales. Terrestrial invertebrates were rarely consumed and consequently relative abundance ranged from 0.02% in France and the

USA to 1.5% in Chile. Variability in relative abundance was highest for vertebrate prey, reaching a maximal relative abundance of 33% in South Africa while being absent in Chile.

According to the PCA (Fig.1.1), the diet of the Portuguese and South African populations was respectively characterized by high relative abundances of ephemeropteran larvae, and eggs of native amphibians. The cluster representing Wales and Chile was characterized by a high occurrence of zooplankton and a very low occurrence of all vertebrate prey items. Except for the eggs of X. laevis, French and American samples share the three most abundant prey categories (i.e. Diptera, Gastropoda, and Zooplankton), which explains the short distance between these samples in the PC space.

Benthic taxa represented 8.54% of the prey categories in South Africa, 6.87% in Wales,

14.85% in Chile, 28.8% in the USA, 55.87% in France, and 93.89% in Portugal (Fig.1.2).

Nektonic taxa represented 39.14% of the prey items in South Africa, 0.14% in Wales, 1.18% in Chile, 1.94% in the USA, 4.20% in France, and 3.01% in Portugal.

Fig.1.2 Occurrence of the main ecological traits among the five populations: terrestrial prey

(black), benthic prey (dark blue), nektonic items (sky blue) and planktonic prey (cyan). Chapter I

Standardized evenness was 0.48 in South Africa, 0.10 in Wales, 0.27 in Chile, 0.42 in the

USA, 0.55 in France, 0.41 in Portugal (Fig.1.3a). The relationship between evenness (J‟) and the number of sampled sites (Ns) followed a logarithmic function (J‟ = 0.086*ln(Ns) + 0.249;

Fig.1.3b). The slope of the curve decreased from one to ten sites, and stabilized at 25 sites.

There was no correlation between trophic niche breadth estimations and the number of stomach contents analyzed per population.

Fig.1.3 Niche breadth calculated for diet data in native and colonized ranges of Xenopus laevis, calculated with the Evenness J‟ (A) and relationship between J‟ and the number of sites NS used in localities (B).

Negative electivity values were observed for most prey items (Fig.1.4). In contrast, zooplankton was preferred (positive electivity) in the Welsh, South African and French populations. The electivity values of the other prey categories were variable, indicating that they were either selected, avoided, or not represented in the sampling.

Fig.1.4 Electivity index for each aquatic prey category consumed in South Africa (brown),

Wales (dark-orange) and France (light-orange). Chapter I

Discussion

In studies focusing on invasive species, both niche conservatism and niche shifts are commonly reported in the literature (Tillberg et al., 2007; Caut et al., 2008; Comte et al.,

2016). Here, we report strong modifications in the realized dietary niche between naturally occurring and introduced or invasive populations of X. laevis. While these differences mainly constituted contractions or expansions of the realized dietary niche, the diet of the Portuguese population represented a shift in the species‟ diet. Individuals from this population were captured in fast flowing streams using electrofishing, while all others frogs were caught in lentic environments. The difference might therefore be attributed to the habitat characteristics, reinforcing the hypothesis that X. laevis is a generalist predator that modifies its diet according to available resources.

Our study provides the first analysis of prey availability for a large number of sites that include both natural and invaded areas. Our findings indicate that X. laevis may expand or shift some dimensions of its trophic niche in novel environments. This result is significant, and has clear consequences for evaluating conditions that favor the invasion of newly introduced populations. Suitable trophic conditions allowing a positive population growth rate may be found in many places where prey abundance is sufficient. A recent macroecological assessment revealed that large areas of suitable climatic conditions were available outside the species‟ native range, and that X. laevis was likely to expand its range in Europe as a result of climate change (Ihlow et al., 2016). The broad global trophic niche of X. laevis and its ability to adapt its diet according to local conditions, contribute to the strong invasive potential of this species, and the high impact it may induce on its environment (see Measey et al., 2016).

The diet of X. laevis has been studied in its native range (Schoonbee et al., 1992; Vogt et al.,

2017), and in different invasive populations (McCoid & Fritts, 1980; Measey, 1998a; Lobos

& Measey, 2002; Faraone et al., 2008; Amaral & Rebelo, 2012; Courant et al., 2014). The

37 first study carried out in the native range was performed in a fish farm (Schoonbee et al.,

1992), which does not necessarily represent the typical diet of native populations. In most studies, including ours, X. laevis was found to predominantly consume relatively small prey items. In Portugal, X. laevis mostly inhabits streams where zooplankton is rare and the main prey items are benthic ephemeropteran larvae. Prey availability in streams was not studied in

Portugal, but collecting such data may help understanding the dietary shift we observed in this population.

In all populations but Chile, predation on amphibians was observed, including X. laevis itself, which represented the most frequent vertebrate taxon in the collective sample. This cannibalistic behavior has been recorded in non-native populations, including in the USA

(McCoid & Fritts, 1980), Wales (Measey, 1998a), Chile (Lobos & Jaksic, 2005), and Italy

(Faraone et al., 2008) but the high frequencies observed in France and South Africa are unprecedented. In the Chilean population, autochthonous amphibians, as well as X. laevis eggs and larvae, were not observed during the sampling period (JM, pers. com.) which may explain their absence in the diet. The small number of stomach contents sampled and sites analyzed for this population may also provide a biased assessment of the putative absence of native amphibians in this area. Predation on amphibians was minor in most populations, except for the native range of X. laevis. This low occurrence of predation on amphibians in other localities may be related to the season during which studies were carried out, and possible changes in the behavior of native amphibians which have co-existed with X. laevis for decades. Our study does not provide any evidence supporting this idea, but ongoing studies should bring new insights into this question. The noteworthy anurophagy reported in this study corroborates the conclusions of Measey et al. (2015) regarding the occurrence of this behavior among pipids. Chapter I

In this study, terrestrial invertebrates represented the least consumed prey class. However, cumulatively, there were as many taxonomic categories among terrestrial, as among aquatic invertebrates. The occurrence of terrestrial prey items did not vary much between populations, suggesting that there were no local specializations for the capture of terrestrial prey. In previous dietary studies of X. laevis, authors concluded that the high portion of terrestrial prey could not exclusively be explained by the capture of terrestrial invertebrates that had fallen into the water (Measey, 1998b) but could represent captures occurring during terrestrial foraging during rainy periods.

From a methodological perspective, sampling effort in each locality was heterogeneous with respect to the sampling period, the number of prospected sites (range: 1-26), and the number of individuals analyzed per population. Consequently, niche breadth could be under-estimated in populations where only few sites were sampled (Wales, Chile, and USA) or where sites were inter-connected (streams in Portugal). According to our results, there is a positive relationship between diet breadth and the number of prospected sites. As no threshold was identified for an optimal number of study sites, we would recommend using as many spatial and temporal replicates as possible for studies aiming at comparing niche breadths. The diet of an individual may be influenced by its size (Schafer et al., 2002), age (Gales, 1982; Rutz et al., 2006), and sex independently of size (Gales, 1982; Göçmen et al., 2011, Van Ngo et al.,

2014). In our study, these data were not available for every population, preventing us from analyzing their effect on diet breadth. In other respects, the unique native population included in our study may not be representative of the diet in the native range. Estimations of niche breadth and prey selectivity may have differed had they been based on a larger sample of native populations. Therefore, we encourage future investigators to consider as many naturally occurring populations as invasive populations in studies aimed at understanding the feeding ecology of invasive animals.

Our results indicate that no prey categories are strongly selected, except for zooplankton in

Wales. This suggests that X. laevis does not usually specializes its diet and hence does not develop a population specific dietary niche. This characteristic may enhance its capacity to establish and spread in novel environments. Potential perturbations of X. laevis on its environment, linked to the large predation on small prey items, are crucial elements of the trophic relationships in aquatic ecosystems that still need to be demonstrated. Our results reflect the diet of X. laevis in invasive populations after decades of colonization and do not necessarily reflect the diet of the species at the moment of introduction. Some invasive species modify their diet during the years, or decades following habitat colonization (e.g.

Tillberg et al., 2007; Gkenas et al., 2016). Comparing the diet of individuals at the core and the edge of a newly colonized area may be an effective approach to investigate the change in dietary composition of X. laevis during the colonization process.

Chapter I

CHAPTER II

Assessing the impact of an invasive population of the

African clawed frog on an amphibian community in

western France. Chapter II

Chapitre II. Evaluation de l’impact d’une population exotique envahissante du Xénope lisse sur une communauté d’amphibiens dans le Centre Ouest de la France.

Courant J, Secondi J, Vollette J, Herrel A, Thirion J-M (under review) Assesing the impact of an invasive population of the African clawed frog on an amphibian community in western France. Amphibia-Reptilia.

Résumé

Les espèces exotiques envahissantes étant une des principales menaces pesant sur le déclin de la biodiversité mondiale, mesurer leurs impacts sur les écosystèmes est un enjeu majeur en biologie de la conservation. Quantifier les impacts d‟une population exotique envahissante représente la première étape dans l‟établissement de plans de conservation efficaces. Dans la présente étude, nous avons appliqué une méthode de suivi par modélisation de l‟occupation des habitats par les amphibiens autochtones pour estimer l‟influence d‟une population exotique envahissante de Xenopus laevis sur la richesse spécifique en amphibiens. Pour analyser l‟impact de cette population exotique envahissante sur les amphibiens autochtones, nous avons pris en compte les caractéristiques des habitats, c‟est-à-dire la structure de la végétation aquatique, la présence d‟autres vertébrés et les paramètres physico-chimiques de l‟eau. Nos résultats ont montrés que la richesse spécifique était significativement négativement corrélée avec l‟abondance en individus de l‟espèce exotique envahissante, ainsi qu‟avec la durée depuis la colonisation, estimée grâce à la distance du site au point d‟introduction de l‟espèce. Nous reportons également une absence de modification de la niche des amphibiens autochtones, en termes d‟habitat, entre les zones colonisées et celles non colonisées. Ceci est susceptible d‟être relié à une trop forte homogénéité des habitats sélectionnés pour l‟étude. Le manque d‟hétérogénéité dans les facteurs abiotiques, leur absence de corrélation avec la richesse spécifique et la forte corrélation entre cette richesse spécifique et l‟abondance du Xénope lisse suggèrent un effet négatif de cette espèce sur la communauté d‟amphibiens autochtones.

Chapter II

Chapter II. Assessing the impact of an invasive population of the African clawed frog on an amphibian community in western France.

Abstract

As invasive species are one of the main threats on global biodiversity, assessing their impacts on ecosystems is a crucial factor in conservation biology. Quantifying these impacts for an invasive population represents the first step in the establishment of efficient management plans. In this study, we applied a method of site-occupancy modeling to estimate the influence of an invasive population of Xenopus laevis in France on the species richness of the amphibian community in the colonized area. To analyze the impact of this invasive species on native amphibians, we took into account the characteristics of the habitat, i.e. the structure of the aquatic vegetation, the presence of other vertebrates, and the physicochemical parameters. Our results show that species richness was significantly negatively correlated with the abundance of the invasive species and the time since colonization, estimated as the distance to the introduction locality of the species in France. We also report an absence of differences in the habitat niche breadth of native amphibians between the colonized and non-colonized areas. This might be related to the homogeneity of the habitats selected for the study. The lack of heterogeneity of the abiotic factors, the absence of a correlation thereof with species richness and the correlation of abundance and time since colonization by Xenopus laevis suggest a negative effect of this species on the amphibian community.

Introduction

Invasive species are among the major causes of the current biodiversity decline (e.g. Alford &

Richards, 1999; Collins & Storfer, 2003; Banci et al., 2013). They are commonly suggested to be abundant and widely distributed in their native areas, and to exhibit several traits like broad environmental tolerance, important genetic variability, short generation time, rapid growth, early sexual maturity, a generalist diet, rapid dispersal, gregariousness and commensalism

(Ricciardi & Rasmussen, 1998; Lockwood, 1999). Invasive populations induce negative effects on native communities by predation, interspecific competition, the spreading of new diseases, or habitat modifications (Ricciardi & Cohen, 2007).

Despite the impacts they have on the conservation of biodiversity, the complexity of the interactions between these mechanisms often discourage research on this issue (Greenlees et al., 2007). Ecological surveys remain necessary to evaluate the impact of invasive species on native ecosystems. Such assessments need to unravel the roles of invasive species and others factors like abiotic parameters on ecosystem modifications (Pagnucco & Ricciardi, 2015).

This distinction is important to discriminate between the causes and consequences of the success of an invasive species and their relations with the alteration of the invaded ecosystems

(Didham et al., 2007). In previous studies, the use of this distinction led to identify invasive species as passengers, drivers, and “back-seat” drivers of changes in ecosystems (for a review, see Bauer, 2012). The assignment of a species to one of these categories is of great help for landscape managers to help them carrying out efficient eradication programs. Indeed, financial and technical resources allocated to biodiversity management are often limited, and providing knowledge about the impacts of an invasive population can help establishing priorities in conservation and eradication plans. Chapter II

In this study, we used abiotic parameters and data related to an invasive population of the African clawed frog, Xenopus laevis (Anura: Pipidae), to assess the role of this invasive species on the modifications of the native amphibian communities in France. Xenopus laevis is widely used in research laboratories across the world (van Sittert & Measey, 2016) for its physiological characteristics, easy maintenance in captivity, the use for pregnancy testing

(Shapiro & Zwarenstein, 1934) and research in developmental biology. Moreover, this species is also noteworthy for being commercialized in the pet trade (Weldon et al., 2007). Several populations have been introduced on four continents after accidental escapes or voluntary introductions (for a review, see Measey et al., 2012). The African clawed frog belongs to the

Pipid family, which is considered as the group where anurophagy is the most frequent amongst amphibians (Measey et al., 2015), suggesting they may impact native species by direct predation. Moreover, X. laevis has been described as a generalist predator in several invasive and native populations (Chapter I). The species is known as an asymptomatic carrier of Ranaviruses (Robert et al., 2007; Robert et al., 2014). The fungus Batrachochytrium dendrobatidis, responsible of the decline of many amphibian populations (Berger et al.,

2016), has long been suspected to be also asymptomatically carried by X. laevis even if no evidence exists in the literature (Fisher & Garner, 2007).

According to Lillo and colleagues (2011), a decrease in reproduction of native amphibians appeared after a few years of colonization by X. laevis in Sicily. However, that study did not take into account neither biotic parameters (except the presence of the invasive frog), nor abiotic parameters. In our study, a method of abundance modeling (Royle &

Nichols, 2003) was used to test the role of abundance, time since colonization, estimated with the distance to the introduction site, and abiotic factors on the species richness of native amphibians. Specifically we tested the hypothesis that the colonization of a site by X. laevis is followed by a negative effect on native amphibians, independent of the abiotic parameters.

Materials & Methods

Data sampling

The African clawed-frog was introduced in France in the village of Bouillé Saint-Paul (Deux-

Sèvres department, western France) after one or several accidental releases in the wild

(Fouquet, 2001; Measey et al., 2012). Since then, the population has expanded and colonized a substantial part of the Maine-et-Loire and Deux-Sèvres departments (Fig.2.1). The area selected for this study is a southeast-northwest transect crossing the colonized range of the species from the introduction site to the colonization front and extending beyond into the non- colonized area in the northwestern part of the transect. This transect was chosen because of the possible correlation between the distance of sites to the introduction site and the duration since those sites were colonized. Despite a lack of quantitative and historical data, especially in the northern part of the population, this correlation is supported by local landscape managers.

Chapter II

Fig.2.1 Distribution of the expanding population of Xenopus laevis in western France. Black dots represent the occurrence localities of the species; yellow dots are the selected study sites and the introduction site is represented by a red star.

All 76 sites used for the monitoring were freshwater ponds (See some example in Appendix

D) with area ranged between 100 and 400 m² and potentially occupied by native amphibians.

We collected habitat parameters that possibly influence the occupancy and/or detectability of species. The water parameters can influence the occupancy by amphibians. Thus, pH (Freda,

1987), temperature (Gulve, 1994), conductivity/resistivity [RES] (Klaver et al., 2013), total dissolved solids [TDS] and turbidity were sampled using a Hannah multi-parameter kit

(Hannah HI-9829-11042) at the end of the first visit on each site. The multi-parameter was calibrated regularly after 10 uses with a quick calibration, and a complete calibration was performed every day before sampling. The structure of the sites was described with three continuous parameters measured in centimeters (water depth, mud depth, and bank height), two categorical parameter describing the maximal and minimal slope of banks in five equal angle categories known to potentially influence the occupancy of anurans (Bosch & Martinez-

Solano, 2003). For the same reasons, we collected data for one categorical parameter describing the main substrate (gravel, stone, sand, mud, leaves) of the pond and two binary parameters describing the presence/absence of an outlet. The potential for annual hydrological regime (permanent or temporary pond) was determined after discussions with the owners of ponds.

The aquatic vegetation can also influence amphibian occupancy (Hartel et al., 2006). Thus, we described this parameter with the percentage of the surface covered by vegetation and with the degree of complexity of this vegetation classified in 14 categories from the simplest (1) to the most complex (14) organization (Lachavanne et al., 1995). The presence-absence of fish was evaluated by collecting data on fish captured during the amphibian data collection with the funnel traps and a net (see below for details). The presence-absence of Myocastor coypus, an invasive mammal that substantially modifies its habitat (Prigioni et al., 2005) was also taken into account. Some habitat parameters possibly change during the year. All these Chapter II parameters were measured the day before the first visit of the amphibian data collection to ensure that the values recorded were as close as possible as the values of the parameters when amphibian data sampling occurred.

The amphibian data were collected to be analyzed with Royle‟s method (Royle & Nichols,

2003). Three visits were performed during three consecutive nights at each pond (from 22 p.m. to 2 a.m.) from the 12th to the 26th of May 2015. Nights with optimal meteorological conditions for amphibian activity (e.g. Bellis, 1962; Cree, 1989; Buchanan, 2006; Saenz et al.,

2006) were favored to perform the visits. During each visit and for all species, the number of adults, larvae and clutches were recorded using standardized protocol duration according to the surface of the site. Five minutes were first dedicated to listening for calling anurans. Then, visual prospection occurred during five minutes for each 100 m² of accessible surface, i.e. where water depth was below one meter and where aquatic or terrestrial vegetation did not prevent it. Then, five minutes for each 100 m² of accessible surface were dedicated to the net exploration. We used a headlamp (Petzl, 200 lumens) to perform the visual prospection and a net with and a 2 mm mesh, 50 cm of net diameter and depth (Roudier, Brie sous Mortagne,

France).

The entire study area is occupied by four urodeles (Lissotriton helveticus, Salamandra salamandra, Triturus cristatus and Triturus marmoratus) and ten anurans (Alytes obstetricians, Pelodytes punctatus, Bufo spinosus, Epidalea calamita, Hyla arborea, Rana dalmatina, Pelophylax lessonae, Pelophylax kl. esculentus and Pelophylax ridibundus) species (Lescure & de Massary, 2012; see Appendix E for illustration). Only adults were considered for the Green frog complex, because these frogs breed later than the other species

(during summer). As they occupy the aquatic habitat boundaries during a large part of the year (Paunovic et al., 2010), their detection was possible even out of their breeding period.

The difficult identification of the species of the green frog group in Western Europe led us to

consider it as a unique taxon, referred hereafter as Pelophylax sp. The presence of adults of X. laevis was verified using submerged funnel traps (60-cm length × 30-cm width × 6-mm mesh diameter) at each site. We used cat food as bait and put floats in the traps to avoid drowning amphibians. The capture effort was adjusted to the surface of ponds, with a pressure of one trap for fifty square meters. The abundance of captured individuals of X. laevis was recorded as well as the distance of the site to the introduction locality.

Data analysis

For each amphibian species/genus, the visits allowed setting up two matrices: a presence- absence and an abundance matrix (3 columns for the visits and 76 rows for the sites). These matrices were used during data analyses. We also built a third matrix, the species richness matrix that indicated the number of native species we detected for each visit and each site.

To model the species richness of native amphibians, we used the N-mixture method (Royle,

2004). One of the most important assumptions of N-mixture methods concerns the need of constancy between the detection probabilities of species. The variability of the detection probability between species was assessed by using the presence-absence matrix of each species to compare the detection probabilities estimated with the single-season site-occupancy null model (MacKenzie et al., 2006) in the Presence software (Hines, 2006). To complete this assessment, a Friedman rank test was performed on the results of the species richness obtained for each visit. In order to model the species richness using the N-mixture method, for each model, one biotic (related to X. laevis, fish, or M. coypus) or abiotic (habitat) parameter was used as a single parameter explaining the species richness of native amphibians. The adjustment of all models was tested with Goodness Of Fit (GOF) tests. Models validated by the GOF tests were compared with one another using their respective corrected Akaike

Information Criterion (AICc). The estimation of species richness according to abundance of X. laevis and distance to the introduction site were provided by the model output. A non- Chapter II parametric Spearman‟s ρ test was also performed between abundance and distance to introduction site to determine the degree of correlation between these parameters.

The effect of X. laevis on the occurrence of each species was assessed by using a recently developed method for analyzing species co-occurrence based on site-occupancy data and using a Bayesian statistical approach (Veech, 2013) in R (R Core Team, 2015) and the cooccur package (Griffith et al., 2016). N-mixture models were applied with the abundance of species exhibiting negative co-occurrences with X. laevis using the same method as the one used to model species richness.

Habitat niche breadth was calculated using the Evenness J’ (Colwell & Futuyma, 1971) with habitat parameters as niche components. The niche overlap between each native species and

X. laevis was calculated with the Morisita‟s overlap index CH (Morisita, 1959). Before the calculation of the indices, a transformation of the habitat parameters into binary variables was necessary. Continuous parameters were transformed in categorical parameters, and then, for each category of this parameter, a binary parameter was created (1 for presence of the category in a site, 0 for its absence). Each categorical parameter underwent the same treatment. Niche breadth was calculated for each native species in the area occupied by X. laevis (J’occ) and in the area unoccupied by the species (J’ctrl). Niche breadth J’ and niche overlap CH indices were calculated according to the following formulas (where, for the

Evenness J’, pi is the proportion of sites occupied by the species for the parameter i and n the total number of parameters; and, for the Morisita‟s overlap index CH, pij and pik are the proportions of sites occupied, respectively, by the species j and k for the parameter i and n the total number of parameters):

2 ∑푛 푝 푝 ′ − ∑(푝푖 ∗ log 푝푖) 푖 푖푗 푖푘 퐽 = 퐶퐻 = 푛 2 푛 2 log 푛 ∑푖 푝푖푗 + ∑푖 푝푖푘

Results

Out of the 76 surveyed ponds used for the survey, 27 of them were located out of the expected distribution range of X. laevis. Moreover, the invasive frog was not detected in ten ponds

(20.41 % of the sites in the occupied area) located in the known colonized area. Two urodele

(Triturus marmoratus and Salamandra salamandra) and two anuran (Alytes obstetricans,

Epidalea calamita) species were not detected during the survey. According to the estimations obtained with the site-occupancy models, the detection probabilities (with the Standard Error,

SE) were high for every species we encountered (Bufo spinosus: p = 0.681 ± 0.045;

Pelophylax sp.: p = 0.840 ± 0.027; Rana dalmatina: p = 0.696 ± 0.043; Hyla arborea: p =

0.841 ± 0.028; Pelodytes punctatus: p = 0.513 ± 0.074; Triturus cristatus: p = 0.673 ± 0.041;

Lissotriton helveticus: p = 0.905 ± 0.020).

Tab.2.1 Summary of the best models obtained with the N-mixture method (Royle, 2004) to estimate species richness. Only models with a better AIC than the null model are listed here.

Resistivity (RES) and total dissolved solids (TDS) were the only abiotic parameters with a better AICc than the null model.

Model AICc ΔAICc AICwgt Mod likelihood Nb Par (-2*LogLike) Abundance 1259.07 0.00 0.2636 1.0000 3 1253.07 Distance 1260.47 1.40 0.1309 0.4966 3 1254.47 RES 1262.46 3.39 0.0484 0.1836 3 1256.46 Presence 1262.73 3.66 0.0423 0.1604 3 1256.73 TDS 1263.13 4.06 0.0346 0.1313 3 1257.13 Null model 1263.39 4.32 0.0304 0.1153 2 1259.39

Chapter II

The highest detection probability was obtained for X. laevis (p = 0.939 ± 0.02). The Friedman rank test, carried out to test the modification of the species richness between visits, did not provide a significant result (Chi² = 0.701; df = 2, p = 0.704). Thus, no changes in species richness were observed between visits. The X. laevis abundance model was the best model of the N-mixture analysis (Tab.2.1), but it did not fit the data better than the distance to introduction site model (ΔAIC = 1.40). According to the Spearman‟s ρ test, X. laevis abundance and distance to introduction site were significantly correlated (S = 126850, p <

0.001, ρ = -0.73). Both models fitted the native amphibian species richness better than the other models (ΔAIC = 1.99). These models were validated by the GOF tests (X. laevis abundance model: Chi² = 30.39; df = 22; p = 0.109; distance to introduction site model: Chi² =

29.45; df = 22; p = 0.133). The X. laevis presence model did not fit the data better than the null model (ΔAIC = 0.66). The X. laevis abundance model and the distance to introduction site model fitted the data better than the presence model (respectively, ΔAIC = 3.66 and ΔAIC =

2.26). The estimations of species richness according to the abundance of X. laevis and the distance to the introduction site are illustrated in Fig.2.2a and Fig.2.2b, respectively. The estimation of the species richness according to the X. laevis presence model, validated by a

GOF test (Chi² = 27.98; df = 22; p = 0.176) resulted in a value of 4.21 ± 0.42 species in absence of X. laevis and a value of 5.07 ± 0.48 species in occupied sites.

Fig.2.2 N-mixture estimates of species richness according to (a) the abundance of Xenopus laevis and (b) the distance to the introduction site. The small black dots represent the data of species richness and the large black dots represent the estimations with the standard errors represented by error bars

According to the co-occurrence analysis, two interactions between species were significantly different from what was expected at random. Xenopus laevis exhibited a significantly negative co-occurrence with B. spinosus (p = 0.009) and T. cristatus (p = 0.020; Tab.2.2). The N- mixture models estimated for B. spinosus and T. cristatus were not validated by the GOF tests, making the interpretation of the modeling of the abundance of these species depending on X. laevis difficult. The niche breadths of the native species in the non-colonized were not different from those of the area colonized by X. laevis (Tab.2.2). All species had a very high niche overlaps with X. laevis (Tab.2.2). Chapter II

Tab.2.2 Effects of Xenopus laevis on the niche position and breadth of native amphibians and its

co-occurrences with these native amphibians. The Evenness J’ has been calculated in the area

occupied by X. laevis J’occ and in the uncolonized area J’ctrl. The modification of Evenness Δ J'

occurring after colonization is also presented. The Morisita‟s niche overlap index CH shows the

overlap between the niches of each native species and X. laevis. The co-occurrence analysis compared

the observed co-occurrence (Obs Cooc) with the expected co-occurrence (Exp Cooc) and determined if

they were significantly different (p-value).

B. spinosus Pelophylax sp. R. dalmatina P. punctatus H. arborea L. helveticus T. cristatus

J’occ 0.214 0.216 0.217 0.212 0.215 0.216 0.212

J’ctrl 0.221 0.221 0.22 0.209 0.219 0.22 0.219 Δ J' 0.007 0.005 0.003 -0.003 0.005 0.003 0.008

CH 0.955 0.975 0.968 0.935 0.965 0.972 0.934 Obs Cooc 17 29 23 11 30 23 13 Exp Cooc 22.6 29.8 24.1 11.3 29.8 24.1 18.0 p-value 0.009 0.444 0.385 0.542 0.655 0.385 0.020

Discussion

In a previous study carried out in the Sicilian invasive population of X. laevis, the authors reported a decrease in the reproductive occurrence of native amphibians correlated with the presence of this invasive species (Lillo et al., 2011). In our study, the results showed that species richness of native amphibians was negatively related to the abundance of X. laevis.

Time since colonization, expressed as the distance to the introduction site, was also negatively related with amphibian species richness. Abundance and time since colonization were correlated, which indicates that X. laevis does not become abundant in ponds immediately after colonization. The longer X. laevis occupied a site, the more abundant it was and the lower the species richness was. One possible cause of this decreased richness could be the negative effects of the invasive African clawed frog on native amphibians. Our results suggested that taking into account the abundance of X. laevis or the distance to introduction improved the modeling of the species richness rather than using the presence of the African clawed frog by itself. This result is not surprising because it implies that the abundance of individuals and the time since colonization provide more detailed information than presence- absence data.

The co-occurrence results showed that two species, B. spinosus and T. cristatus, co-occurred negatively with X. laevis. This result is different from the one obtained in Sicily, where an impact was detected for anurans belonging to the Ranidae, Hylidae and Discoglossidae (Lillo et al., 2011). The method used by Lillo and colleagues (2011) was not the same as in our study, which prevents us to perform a quantitative comparison. However, the differences in the effects observed in France and Sicily might be related to the fact that the impacts of a given invasive species are commonly more pronounced in insular than in continental areas

(e.g. D‟Antonio & Dudley, 1995; Gimeno et al., 2006). Chapter II

The niche breadth of the native amphibians was not modified in the area colonized by X. laevis. Moreover, the niche overlap between the native amphibians and X. laevis was very high for every species. This could indicate that the niches of native amphibians were not reduced when they share their habitat with the invasive frog. However, the constancy of niche breadths and overlaps could be due to the relative similarity between the study sites. It is likely that we did not sample a large enough part of the realized niche of each species to detect potential changes on niche breadth or position. An alternative possibility would be our failure to efficiently sample habitat data allowing detecting changes in niche breadths and overlaps. To correct these potential weaknesses, a more important diversity of sites should be sampled to explore the complete realized niche of native and invasive species. It would help to improve the assessment of the role of the invasive species and the abiotic parameters on the decline of a native community. However, even if assessing how the abiotic factors influence the native amphibian community with our results is impossible, they clearly demonstrate that

X. laevis has a key role, as driver or back-seat driver, in the modifications of the amphibian community.

Several mechanisms could underlie the impact of X. laevis on native amphibians in western

France as recorded in this study. Predation on amphibians has been recorded among several families of anurans, but the African clawed frog belongs to the family where the frequency of anurophagy is the most important (Measey et al., 2015). The larval stages of B. spinosus are toxic for many predators, but occasional consumption of larval Bufonids has been recorded in the introduced population of Sicily (Faraone et al., 2008). The capture of adult newts (L. helveticus) by X. laevis, as well as the consumption of anuran tadpoles, including X. laevis itself, has recently been reported (Chapter I). The extensive consumption of larval amphibians might be underestimated because of the quick decomposition by gastric juice (e.g. Measey,

1998; Lillo et al., 2011). Thus, the existing literature suggests that predation on anuran by X. laevis might play a role in its impacts on the native amphibian community.

Interspecific competition is another widespread mechanism that may drive the impacts of invasive species on native communities. This has been demonstrated for several invasive plants that monopolize soil and light resources (Lockwood et al., 2013). Competition also occurs in the case of invasive animals able to capture a wide variety of prey items, such as the common rat, Rattus rattus (Caut et al., 2008). Xenopus laevis consumes small invertebrates like dipteran larvae and zooplankton in large proportions (Measey, 1998; Chapter I). These prey items are essential in the trophic cascade of aquatic ecosystems like ponds (Brett &

Goldman, 1996). A perturbation induced on zooplankton might result in effects on other predators of this class of prey, such as the native urodeles. The African clawed frog is also known as a carrier of the chytrid fungus, B. dendrobatidis, which is held responsible of the decline of many amphibian populations on several continents (e.g. Berger et al., 1998; Bosch et al., 2001; Olson et al., 2013). The African clawed frog is also known as an asymptomatic carrier of ranaviruses (Robert et al., 2014) and has been detected with a high prevalence of these viruses in an invasive population (Soto-Azat et al., 2016). These pathogens are responsible of population declines in amphibians (Daszak et al., 2003). Even if the relation between the appearance of the disease and the expansion of an invasive species is hard to prove once the colonization of an area has already occurred (Kumschick et al., 2017), some special attention to this issue is necessary in the context of conservation programs.

From a methodological point of view, the detection probabilities estimated for each species indicate that the sampling design allowed a successful detection of each species after three visits. The conditions to apply the N-mixture analysis – equal detection probability among species and visits – were validated. However, our sampling design shows spatial and temporal weaknesses. The study only occurred during a single season, which prevented us to take into Chapter II account time-related parameters like the ageing of ponds and the effect of weather on the occurrence of each species. Modifications of ponds over time likely result in changes in the amphibian community. Thus, the age of ponds might influence the abundance of X. laevis, and may be another driver of the ecological change of the amphibian community. Moreover, for a given pond, the potential for occupancy by each amphibian species changes during the year, depending on its use (cattle) and on climatic and environmental parameters (See

Appendix F for an illustration). We only took into account abiotic parameters describing the ponds and their immediate surroundings. However, the nature of the landscape and its connectivity is an important factor of occurrence and abundance for amphibian populations

(Rothermel, 2004; Cushman, 2006; Ribeiro et al., 2011). Another weakness of this study concerns the possible changes in the value of abiotic parameters during the year and the period of reproduction of amphibians (e.g. water depth, aquatic vegetation) or through the day and night (e.g. temperature). Determining at which moment the value of each factor was the most determinant for amphibian activity was not possible for this study. It is likely that the evolution of a parameter through time is more important than its value at a given moment.

Thus, the influence of abiotic parameters might be underestimated. Finally, the sampling design of this study prevented us from distinguishing the effects of two correlated parameters, the invaders‟ abundance and time since colonization. To allow this discrimination, the effect of the abundance of an invasive species on an ecosystem should be evaluated using a sample design where time since colonization is rigorously known. The effect of time since colonization should be assessed by comparing the native species richness between a large enough number of sites colonized at several dates with an equal abundance of invasive individuals in every site.

Chapter II

Second part

Introduction

Chapter III

In the first two chapters, we analyzed the direct interactions of the African clawed frog with its environment in the French invasive population. The following chapters use a different approach, and aim at providing findings about the changes in resource allocation in the same population, currently facing an expansion. These findings may help assessing the potential for long range expansions of this species, and, as a consequence, quantifying its long term impacts.

The phenomena occurring during range expansions have been paid a great interest during the last decades because of their importance for the setting up of management plans for the conservation of species facing climate change or invasive species, or for reintroduction programs (Mack et al., 2000; Thomas et al., 2004; Roman & Darling, 2007; Shine 2012). As mentioned in the introduction of this thesis, the expansion rate of a population is related with its growth rate, the density-dependence and the dispersal abilities of the individuals that constitute the population. The trade-offs between the life-history traits summarized by these population parameters drive the probabilities for individuals to disperse, reproduce and survive, respectively (Burton et al., 2010).

To apply this approach, we used the expanding population of X. laevis in France to study the changes in resource allocation to the life history traits. This species is known to exhibit high reproductive capacities in invasive contexts (McCoid & Fritts, 1989), but understanding how these capacities change during range expansion remain unknown. Despite its predominantly aquatic lifestyle, it has been able to colonize large areas in France and other countries. In this regard, understanding the relative occurrence of overland movements in this species remains poorly documented (Measey, 2016). In this regard, and because this species is among the amphibians that exhibit the highest ecological impacts (Kumschick et al., 2017), studying resource allocation to life history traits appears as an important element in conversation biology. Chapter III

We followed the predictions of Burton and colleagues (2010) and the general pattern observed in empirical studies to put forward hypotheses for each life history traits. In the following chapters, resource allocation to life history traits has been studied following the sampling design described in Fig.Int.2. Reproduction has been compared between the range core and the range edge in Chapter III. Resource allocation to dispersal has been studied through the endurance performance in Chapter IV, and through a field survey using a capture-mark- recapture method, and experiments on movements in Chapter V. Finally, in Chapter VI, we used the same field survey as in Chapter V to determine how the population dynamics change during the range expansion of X. laevis in France.

Chapter III

CHAPTER III

Resources allocated to reproduction decrease at the range edge of an expanding population of an invasive amphibian. Chapter III

Les resources allouées à la reproduction diminuent sur le front de colonization d’une population en expansion chez un amphibian exotique envahissant.

Courant J, Secondi J, Bereiziat V, Herrel A (2017) Resources allocated to reproduction decrease at the range edge of an expanding population of an invasive amphibian. Biological Journal of the Linnean Society. https://doi.org/10.1093/biolinnean/blx048 [Appendix G]

Résumé

Prédire l‟importance et la nature des changements dans l‟aire de répartition d‟une espèce devient de plus en plus important à mesure qu‟augmentent le nombre d‟espèces faisant face à des modifications de leurs habitats, ou à des introductions hors de leurs aires de répartitions naturelles. De par la rencontre avec de nouvelles conditions environnementales, l‟investissement d‟un organisme dans les traits d‟histoire de vie est supposé changer pendant les modifications d‟aires de répartitions. Alors que les simulations prévoient une augmentation de l‟allocation des ressources à la dispersion et la reproduction sur le front de colonisation, nous proposons que l‟allocation à la reproduction devraient diminuer sur le front de colonisation, à cause de compromis évolutifs dans l‟allocation des ressources. Nous avons étudié l‟investissement dans la reproduction chez un amphibien exotique envahissant, Xenopus laevis, et nous avons mesuré l‟allocation à la reproduction dans trois groupes de populations distribuées depuis le centre de l‟aire de répartition vers le front de colonisation de l‟espèce en France. L‟allocation des ressources a été estimée avec un indice de masse corrigée des gonades, le Scaled Mass Index, pour les deux sexes pendant la période de reproduction de l‟espèce. The degré de ressources allouées à la reproduction était plus faible sur les marges de l‟aire de répartition comparé à son centre. Ceci pourrait être le résultat de changements dans les compromis évolutifs entre les traits d‟histoire de vie. Ce patron pourrait être expliqué par la compétition interspécifique ou par un investissement plus prononcé dans la dispersion sur le front de colonisation.

Chapter III

Resources allocated to reproduction decrease at the range edge of an expanding population of an invasive amphibian.

Society. https://doi.org/10.1093/biolinnean/blx048 [Appendix G]

Abstract

Predicting the magnitude and nature of changes in a species' range is becoming ever more important as an increasing number of species are faced with habitat changes, or are introduced to areas outside of the species‟ native range. An organism's investment in life history traits is expected to change during range shifts or range expansion because populations encounter new ecological conditions. While simulation studies predict that dispersal and reproductive allocation should increase at the range edge, we suggest that reproductive allocation might decrease at the range edge due to energy allocation trade-offs. We studied the reproductive investment of an invasive amphibian, Xenopus laevis, and measured reproductive allocation in three clusters of populations distributed from the center to the edge of the colonized range of X. laevis in France. Resource allocation was estimated with the Scaled Mass Index of gonads for of sexes during the local period of reproduction of the species. The level of resources allocated to reproduction was lower at the periphery of the colonized range compared to the center and may be the result of changes in trade-offs between life history traits. Such a pattern could be explained by interspecific competition or enhanced investment in dispersal capacity.

Introduction

Shifts in the distribution of a species occur when range edges track environmental changes, when propagules settle beyond an ecological barrier after a discrete introduction event, or when evolutionary processes modify the species‟ niche and render novel environmental conditions suitable. This issue has been paid renewed attention because of the increasing awareness about the scale of global changes (Parmesan & Yohe, 2003), including human activities that translocate individuals to new areas (Channell & Lomolino, 2000). In this regard, the massive increase in global transport and trade is believed to be responsible for the fast rise in the number of invasive species at the global scale (Lockwood et al., 2013; Herrel

& van der Meijden, 2014). Analyzing the process of range expansion is critical to understand and predict the dynamics of a species. This knowledge is also important to guide the management of practitioners who seek to enhance the expansion of reintroduced populations, or limit the spread of unwanted aliens (e.g. Roman & Darling, 2007; Shine, 2012; Brown et al., 2013).

Range expansion typically occurs after the introduction and settlement of a propagule in a novel environment. It is largely driven by population growth, dispersal, and density dependence (Fisher, 1937; Skellam, 1951; Burton et al., 2010). These parameters summarize the trade-offs in the life history traits that influence the probability for an individual to reproduce, disperse, and survive (Burton et al., 2010). Simulation and empirical studies have demonstrated that allocation of resources to dispersal at the range edge is enhanced during range expansions (e.g. Thomas et al., 2001; Travis & Dytham, 2002; Simmons & Thomas,

2004; Hughes et al., 2003; Phillips et al., 2008; Léotard et al., 2009; Phillips et al., 2010;

Bartoń et al., 2012; Henry et al., 2013; Hudson et al., 2016). Changing dispersal allocation affects the trade-offs with other life history traits that also require energy or resources

(Ferrière & Le Gaillard, 2001). Furthermore, Burton and colleagues (2010) predicted that Chapter III dispersal and reproductive allocation should increase whereas resource allocation to competitive ability should be reduced at the range edge. The rationale is that long dispersers have a higher probability to reach a new suitable location at the edge where density-dependent responses and intra-specific competition are low (Brown et al., 2013). At the range core, dispersers would mostly reach already colonized areas (Baguette & Van Dyck, 2007) so that in core areas selection should favor individuals with a high competitive ability (Burton et al.,

2010). In contrast with these predictions, the association of high dispersal capacity and low reproduction has been observed (Hughes et al., 2003). It is therefore important to test whether resource allocation to reproduction is reduced or increased at the range edge of expanding populations.

We studied the allocation of resources to reproduction in an expanding population of the

African clawed frog, Xenopus laevis (Daudin, 1802) (Anura: Pipidae) in Western France, where it has been introduced. This species has been used worldwide in laboratories for pregnancy tests and as a model in developmental biology and anatomy (Weldon, De Villiers

& Du Preez, 2007; van Sittert & Measey, 2016). It has been intentionally or unintentionally introduced in several countries across the world where it is considered as invasive (Measey et al., 2012). A study based on Species Distribution Models (SDMs) suggests that the expansion of the French population is likely to be important during the next decades (Ihlow et al., 2016).

In another recent study, an increased allocation to dispersal has been suggested through the observation of an increase in hind limb morphology and stamina at the range edge of the same population (Chapter IV). Thus, we hypothesize that at the range edge individuals will allocate fewer resources to reproduction than at the range core. We test this hypothesis by estimating the relative mass of the reproductive organ of males and females with a body condition index.

Material & Methods

Study area and sampling

The African clawed frog was introduced in Western France during the late 1980s, and its estimated colonized area was 207 km² in 2012 (Measey et al., 2012). Contrary to the

Mediterranean like climate of its native range, the population introduced in France occupies an area where the climate is defined as oceanic altered (Joly et al., 2010), with a relatively high annual mean temperature (12.5°C), and cumulated annual precipitations reaching 800 to

900 mm, mainly during winter (Joly et al., 2010). We sampled ponds with an area of less than

500 m² and a maximum depth between 1 and 3 meters. Frogs were captured with fykes (60 cm length x 30 cm width, 6 mm mesh diameter). Capture effort was fixed at 1 trap per 50 m² of water, with a distribution of traps as regular as possible inside the pond, during three consecutive nights, and without repeating this process more than once in the same pond.

Every morning, individuals captured in fykes were euthanized with a lethal injection of sodium pentobarbital and dissected just after death. The research permit and the authorization for killing invasive frogs were provided by the Préfet of the Deux-Sèvres department.

Determination of the optimal period to study reproduction

In California, the period of reproduction of X. laevis occurs from January to November, with a two to three months peak during spring (McCoid & Fritts, 1989). For the French population, the optimal time frame to estimate reproductive allocation was determined by sampling the gonad mass and the proportion of gravid females from March to October 2014 (no individuals were caught during trapping sessions from November to February). We captured and analyzed a total of 408 female frogs to determine the breeding period of this population, with monthly captures in 6 ponds from the range edge, core and intermediate areas (two in each area), except in October when only two successful capture occurred at the range core and the intermediate area. New ponds were used every month. The optimal time frame was then Chapter III defined as the period when the proportion of gravid females and the relative mass of reproductive organs (estimated with the Scaled Mass Index, see below) was increasing and when the females are most likely to start breeding and have to laid many eggs. We then could assess the spatial variation of allocation to reproduction on the next spring, in 2015.

Estimation of the resource allocation to reproduction

The present study was carried out on three clusters of five ponds each, distributed across a part of the colonised range (Fig.3.1). The clusters were located at the core area close to the introduction site of the population, the range edge (18km from the introduction site) and in an intermediate zone between both areas (11km from the introduction site), respectively. The time since colonization at the center of the range is close to 30 years (Fouquet, 2001). At the range edge, the study sites were colonized less than 3 years ago, according to local landscape managers and unpublished data. This area was selected because it was the most accurately identified colonization axis during our field studies and because no barrier for dispersion was identified in the landscape. In 2015, 731 frogs were captured to study the resource allocation to reproduction, with a 3 ponds per week frequency (one pond in each cluster). We captured

82, 102 and 120 females at the range core, the intermediate area and the range edge, respectively. For males, 186, 103 and 138 individuals were captured at the range core, the intermediate area, and the range edge, respectively. Data sampling for the reproductive allocation study occurred from the 4th of April to the 13th of May 2015.

Fig.3.1 Location of the three pond clusters of Xenopus laevis sampled to analyze the variation in resource allocation to reproduction. Occurrence data for X. laevis was provided by the Société

Herpétologique de France. Chapter III

Morphological and anatomical measures

Individuals were measured and sexed before dissection during 2014 and 2015 data samplings.

Males were recognized by the presence of reproductive callosities on their forelimbs and by the flat shape of their cloaca. Females were identified by the dilated shape of the cloacae and the absence of reproductive callosities on their forelimbs. The combination of these criteria avoided misidentification of sex for adults. However, for small individuals (snout-vent length

< 50 mm for males, and < 70 mm for females), sex could not be determined unambiguously based on external characters. Individuals measuring less than these threshold values were thus not considered as sexually mature and not included in our analysis. Snout-vent length (SVL) was measured with a digital caliper (Mitutoyo Absolute IP67 – precision 0.01mm). Gonads, the adipose tissues associated to them and body mass were weighed using a digital balance

(XL 2PP – precision 0.01g). In order to avoid any measurement bias due to the mass of ingested prey items, the stomach content was removed from each individual before measuring body mass. Immature oocytes were detected by the aspect of ovaries. Ovarian stages are numbered from I to VI. Oocytes at stages I-III are completely plain, whereas oocytes at stages

IV-VI are bipolar. Therefore, we considered females as being reproductively active from stage IV onwards (Rasar & Hammes, 2006) and we did not use females with oocytes in earlier stages. Reproductive effort is one of the main components of the reproductive strategy of amphibians (Duellman & Trueb, 1986) and it allows a measurement of the costs of reproduction for many organisms (Stearns, 1992), especially for those that do not provide parental care. The role of adipose tissue associated to gonads consists in providing resources to future reproduction events during the reproductive season (Girish & Saidapur, 2000) and is thus an important component of reproductive investment.

Statistical analyses

To estimate allocation to reproduction, we adjusted the mass of reproductive organs (gonads + adipose tissue) by using a body condition index. Most indices found in the literature are based on an ordinary least square (OLS) regression between body size (SVL) and body mass.

However, those indices do not deal with overdispersion and heteroscedasticity (Peig & Green,

2010). Furthermore, it has been recommended not to use OLS based indices for studies comparing different populations or sites (Jakob, Marshall & Uetz, 1996). We used the Scaled

Mass Index (SMI) which does not present these limitations (Peig & Green, 2009). The SMI has already been tested on an anuran, Lithobates catesbeianus, and provided a good fit to the data (MacCracken & Stebbings, 2012). This index, noted 푀 푖, is calculated according to the following formula:

푏푆푀퐴 퐿휃 푀 푖 = 푀푖. 퐿푖

Where 푀푖 is the mass of the individual i, or the tested organ, Li the SVL of the individual, 퐿휃 the mean SVL of the sample (for details, see Peig & Green, 2009). The exponent bSMA is the slope of the standard major axis (SMA) regression of M on L. SMI values were estimated based on a 퐿휃 and bSMA calculated separately for each cluster and each sex. This index adjusts the body mass of individuals to the mass they would have at the length Lθ. The corrected mass, calculated with the SMI for reproductive organ, gonads, adipose tissues and body mass will be referred in the following sections as SMIorgan, SMIgonad, SMIadipose and SMIbody, respectively. To determine the best period to study reproductive allocation and to assess spatial variations, we used the data collected in 2014. We tested the variation of the SMIorgan between months using an analysis of variance (ANOVA). Pairwise differences in SMIorgan between consecutive months were tested using t-tests with an adjustment of α by applying a

Bonferroni correction. Chapter III

With the data collected in 2015, we calculated the proportion of gravid females in each cluster

(number of gravid females/ total number of females). Using the same dataset, we investigated the spatial variation in resource allocation between clusters by comparing the SMIorgan in clusters and by using a generalized linear mixed model for each sex with cluster type and month as fixed effects and capture site nested within cluster as a random effect. We took into account the effect of body condition by adding the SMIbody in the linear mixed models as a fixed effect. We also provided the mean values for SMIgonad, SMIadipose and SMIorgan for each cluster in a table. All analyses were carried out using R version 3.0.1 (R Core Team, 2013) and the packages ggplot2 for graphics and nlme (Pinheiro et al., 2013) for the linear mixed models.

Fig.3.2 Mean values (± Standard Error) of the SMIorgan of females Xenopus laevis captured in

2014. The number of individuals (N) and the proportion of gravid females (FGravid) captured each month is mentioned close to the dot of each month.

Chapter III

Results

Determination of the breeding season

In 2014, The SMIorgan was not constant during the period of activity of the species (ANOVA, relative ovary mass: F7,335= 2.61, p<0.05). It was significantly higher in March compared to

April (Welch Two Sample t-test: t21.28= -6.8993, p< 0.001), but the capture success and the resulting sample size in March were low. The SMIorgan decreased down to the baseline level from July to October (Fig.3.2), with a significant increase from April to May (Welch Two

Sample t-test: t172.45= 2.66, p< 0.01), a significant decrease from June to July (Welch Two

Sample t-test: t52.44= 2.91, p< 0.01) and from September to October (Welch Two Sample t- test: t7.12= -11.86, p< 0.0001). There was no significant change between March and April

(Welch Two Sample t-test: t16.14= 0.06, p= 0.95), May and June (Welch Two Sample t-test: t39.45= -1.43, p= 0.16), July and August (Welch Two Sample t-test: t111.15= 0.72, p= 0.48),

August and September (Welch Two Sample t-test: t18.60= 1.44; p= 0.17).

Tab.3.1 Results of the linear mixed models performed to test the spatial variation in reproductive allocation across the colonized range of Xenopus laevis in France.

Effects: Estimate Std.Erro Df t- p- r value value

Females Intercept 0.853 0.849 245 1.004 0.316

Cluster Core VS Edge -2.359 0.923 23 -2.555 0.018

Cluster Core VS Inter -2.427 0.989 23 -2.454 0.022

Month -0.286 0.769 245 -0.272 0.710

SMIbody 0.106 0.009 24 11.98 <0.00 9 1

Random: Cluster/Pond 1.235 2.844 - - -

Males Intercept 0.142 0.010 377 1.364 0.173

Cluster Core VS Edge -0.025 0.010 24 -2.384 0.025

Cluster Core VS Inter -0.008 0.012 24 -0.701 0.490

Month -0.003 0.006 377 -0.447 0.655

SMIbody 0.003 <0.001 377 14.38 <0.00 0 1

Random: 0.017 0.035 - - - Cluster/Pond

Chapter III

Spatial variation in resource allocation

In 2015, the proportion of gravid females was 66.26 % for the core, 78.43 % for the intermediate, and 51.67 % for the edge cluster respectively. According to linear mixed models, allocation to reproduction for females decreased from the range core to the intermediate (SMIorgan(Core) – SMIorgan(Inter) = 1.46) and edge (SMIorgan(Core) –

SMIorgan(Edge) = 0.92) ranges (Fig.3.3 and Tab.3.1). In males, SMIorgan was not significantly higher (Tab.3.1) in the core cluster than in the intermediate cluster (SMIorgan(Core) –

SMIorgan(Inter) = 0.34) and significantly higher (Tab.3.1) at the core than at the range edge

(SMIorgan(Core) – SMIorgan(Edge) = 0.37). The decrease in SMIorgan mainly concerned the gonad relative mass for females whereas, for males, both testicles and adipose tissue masses were reduced at the range edge (Tab.3.2).

Tab.3.2 Mean values and Standard Errors obtained for each sex for the SMI of gonads, adipose tissue and entire reproductive organ in the three clusters.

Core Intermediate Edge Females

SMIgonad 5.36±0.56 3.80±0.29 4.09±0.46

SMIadipose 1.02±0.12 0.88±0.10 1.71±0.15

SMIorgan 6.14±0.56 4.67±0.25 5.22±0.49

Males

SMIgonad 0.11±0.004 0.11±0.004 0.08±0.003

SMIadipose 1.22±0.06 0.88±0.05 0.89±0.05

SMIorgan 1.33±0.06 0.99±0.05 0.96±0.06

Fig.3.3 Reproductive investment of Xenopus laevis in the core, intermediate, and edge clusters as represented by the SMIorgan (mean ± SE) of females (left) and males (right).

Chapter III

Discussion

In this study, we observed a modification of the reproductive investment of X. laevis in

Western France. This modification consists in a decrease of the relative mass of reproductive organs during the first part of the peak of reproduction from the core of the colonized range to its periphery.

Determination of the breeding season

According to our results based on the data collected in 2014, the proportion of sexually active females varied between March and October in the French population. Reproductive activity peaked for three months between April and June. In South Africa, the breeding period lasts three to five months (Deuchar, 1975). The maximal breeding activity lasts only two to three months in California, but reproduction can occur from January to November (McCoid &

Fritts, 1989). The duration of the optimal breeding period in France and South Africa cannot be strictly compared because Deuchar (1975) only mentioned the breeding period and did not define the optimal breeding period in itself. In California, water temperature in ponds reaches

20°C each month of the year. Such a high temperature regime could favor year-round reproduction (McCoid & Fritts, 1989). In France, a drop in temperature could explain the sharp decrease in the proportion of gravid females and the relative gonad mass at the beginning of autumn when air temperature usually remains below 20°C. Temperature, among others, affects the variation of food abundance, which could be a crucial factor for the sexual activity of the species (Holland & Dumont, 1975; Girish & Saidapur, 2000). In 2015, to study reproductive allocation, we only considered April and May. We excluded June because most females captured at this moment are likely to have already laid some eggs during the previous months.

Spatial variation in resource allocation

In 2015, we found that at the beginning of the optimal breeding period, the proportion of gravid females was lower at the range edge, but the effect was rather weak. A lower frequency of reproductive females has already been observed in an invasive amphibian during range expansion (Hudson et al., 2015). However, these authors were not able to unambiguously link this pattern to a lower investment in reproduction at the range edge because their data were not initially collected for that purpose. Also in our study we cannot provide an unambiguous conclusion as the non-gravid females that we caught may have bred earlier in spring or may reproduce later in the summer. Females from the core area may also be able to restart ovary development earlier than females from the range edge but laboratory studies remain necessary to test this hypothesis.

The masses obtained with the SMI appeared much more informative, with significantly lower values at the range edge compared to the core area for both sexes. These modifications concerned the testicles and adipose tissues for males, both reduced at the range edge. In anurans, this tissue is crucial to allow spermatogenesis in males and the priming of sexual behavior (Iela et al., 1979; Grafe et al., 1992). At the opposite, no changes in adipose tissue mass have been detected for females. Its extent is minimal during the breeding season (Iela et al., 1979) and there is a negative correlation between adipose tissue and ovary mass (Girish &

Saidapur, 2000). The lack of differences in adipose tissue mass for females between the range edge and the core area could be explained by the time of the study, i.e. the first two months of the peak of reproductive activity. For both sexes, stronger differences are expected when the development of the adipose tissue is maximal.

The results of our study are corroborated by the increased allocation to dispersal observed in the same invasive population of X. laevis (Chapter IV). A study using individual-based models suggested that resource allocation may change depending on the presence, the Chapter III reproductive rate, and the competitive ability of native competitors (Burton et al., 2010). At low density of competitors at the range margin, resource allocation to competitive ability is minimal while resources allocated to dispersal and reproduction are maximal. At higher density of competitors at the range core, allocation to reproduction decreases when competitors exhibit high reproductive rate and competitive ability. This trait has been studied most commonly in invasive plants (Strauss et al., 2002) and is defined as the effect of an individual on resources and its response to resource deficiency (Goldberg, 1990). It influences the survival probability of individuals as a lower investment in competitive ability leads to a reduced survival in high density conditions (Burton et al., 2010).

In the colonized area, the aquatic habitats used by X. laevis are occupied by several native amphibian species that breed during the same period of the year (Lescure & de Massary,

2012). During field work, we observed native amphibians like Pelophylax sp. in high densities in every pond. Other species (e.g. Rana dalmatina, Lissotriton helveticus, Triturus cristatus) were often observed at lower densities. Unfortunately, even if impacts of X. laevis on other amphibians have already been shown (Lillo et al., 2011), no study has assessed the competition between invasive and native amphibians. Competition through trophic resource access is likely to occur (Courant et al., pers. obs.). Further investigations testing the impact on native amphibians may provide some clarification about this issue.

Because females lay eggs in groups of one, two or three eggs (Crump, 2015) it is not possible to tell whether a female has already laid some eggs. We limited this effect by analyzing data collected during the first two months of the peak of reproductive activity. It is also important to notice that every female is subject to this effect irrespective of the locality where they are captured. Consequently, this effect should not drastically alter the conclusions of this study, but it should be taken into account in future studies using the results from the present study

(i.e. care should be taken in interpreting these data as absolute values of reproductive

investment). The pattern of variation in reproductive organ mass that we observed is primarily expected to reflect changes in the allocation of resources to different life history traits.

However, the phenomenon called gene surfing, i. e. the fact that mutations occurring on the range edge travel with the wave front and induce mutant populations (Edmonds et al., 2004), could provide an alternative explanation. This process may sort a different pool of alleles in different directions as propagules are loosely connected at the margins of an expanding range

(Excoffier et al., 2009). Genetic analyses may help to solve this issue. Sampling transects pointing away from the range core in different directions would certainly also help to provide more robust conclusions. Finally, common garden studies on resource allocation in expanding populations have already demonstrated that heritability could have an effect on the observed pattern for a life history trait (Brown et al., 2014). The African clawed frog is an interesting model to address this issue because the ecological conditions (climate, communities) encountered in its native range and the colonized areas differ. This species, by offering intraspecific comparisons across populations introduced independently in different countries may bring insights into the processes of range expansion and invasion. Chapter III

CHAPTER IV

Differences in mobility at the range edge of an expanding invasive population of Xenopus laevis in the west of France Chapter IV

Différences dans la mobilité sur le front de colonization d’une population en expansion de Xenopus laevis dans l’ouest de la France.

Louppe V, Courant J, Herrel A (2017) Differences in mobility at the range edge of an expanding invasive population of Xenopus laevis in the west of France. Journal of Experimental Biology. 220, 278-283. doi:10.1242/jeb.146589 [Appendix H]

Résumé

Les modèles théoriques prévoient qu‟une mobilité accrue pour les individus situés sur le front de colonisation devrait être favorisée par rapport aux individus situés dans le centre de l‟aire de répartition des populations en expansion. Malgré le fait que les données empiriques sur l‟évolution des performances locomotrices sur les fronts de colonisation sont toujours rares, les données sur le crapaud buffle supportent ces modèles théoriques. Cependant, déterminer si ce phénomène est généralisable aux autres espèces en expansion demeure inconnu. Dans cette étude, nous fournissons des résultats sur l‟endurance locomotrice et la morphologie des membres chez des individus capturés dans deux sites : au centre et au front de colonisation d‟une population en expansion du Xénope lisse, Xenopus laevis en France, où il a été introduit il y a plus de trente ans. Des données de morphologie provenant de deux sites additionnels ont été ajoutées aux résultats afin de tester si les différences observées pour le premier jeu de données peuvent être généralisées dans la population introduite en France. L‟espèce arborant un fort dimorphisme sexuel, nous avons également testé les différences entre sexes pour les performances locomotrices et la morphologie. Nos résultats ont montrés un dimorphisme sexuel significatif dans l‟endurance et la morphologie, avec des mâles présentant de plus longs membres postérieurs et une meilleure endurance que les femelles. De plus, en accord avec les modèles prédictifs, les individus du front de colonisation montraient une meilleure endurance. Cette différence dans les performances locomotrices est susceptible d‟être due aux différences dans la longueur relative des membres postérieurs observés dans les sites sélectionnés. Nos résultats ont des implications en biologie de la conservation, car les modifications observées pourraient conduire à une accélération de la vitesse d‟expansion de cette espèce exotique envahissante en France.

Chapter IV

Differences in mobility at the range edge of an expanding invasive population of

Xenopus laevis in the west of France

Abstract

Theoretical models predict that, at the range edge of expanding populations, individuals with increased mobility should be favored relative to individuals at the center of the range. Despite the fact that empirical evidence for the evolution of locomotor performance at the range edge is rare, data on cane toads support this model. However, whether this can be generalized to other species remains largely unknown. Here, we provide data on locomotor stamina and limb morphology in individuals from two sites: one from the center and one from the periphery of an expanding population of the clawed frog Xenopus laevis in France where it was introduced about 30 years ago. Additionally, we provide data on the morphology of frogs from two additional sites to test whether the observed differences can be generalized across the range of this species in France. Given the known sexual size dimorphism in this species, we also test for differences between the sexes in locomotor performance and morphology. Our results show significant sexual dimorphism in stamina and morphology, with males having longer legs and greater stamina than females. Moreover, in accordance with the predictions from theoretical models, individuals from the range edge had a greater stamina. This difference in locomotor performance is likely to be driven by the significantly longer limb segments observed in animals in both sites sampled in different areas along the range edge. Our data have implications for conservation because spatial sorting on the range edge may lead to an accelerated increase in the spread of this invasive species in France.

Introduction

Dispersal, typically defined as a permanent movement away from a site of birth and/or reproduction (Clobert et al., 2009), may be due to discrete and repeated introduction of propagules beyond ecological barriers. Alternatively, evolutionary mechanisms that make new ecological niches accessible may also spur dispersal (Wilson et al., 2009). Dispersal is dependent on many factors including an organism‟s phenotype, sex, age, reproductive output, the intensity of competition and environmental conditions (Stevens et al., 2010). The synergistic influence of these parameters may result in the spatial differentiation of expanding populations (Shine et al., 2011). Spatial sorting and lower population density have been documented at the invasion front of expanding populations (e.g. Phillips et al., 2010).

Theoretical models based on these observations have subsequently predicted increased dispersal and reproductive rates along the edge of the expansion range (Hallatschek & Nelson,

2007; Excoffier & Ray, 2008; Excoffier, 2009; Burton et al., 2010; Travis et al., 2010).

Lower predation pressure by specialist predators and reduced competition (Burton et al.,

2010; Phillips et al., 2010; Brown et al., 2013), in addition to an increased risk of kin competition resulting from low population density (Kubisch et al., 2013), may also encourage an increased dispersal rate at the colonization front. However, processes such as mutation surfing (Travis et al., 2010), Allee effects (Travis & Dytham, 2002) or allocation trade-offs

(Bishop & Peterson, 2006; Fronhofer & Altermatt, 2015) may prevent this from happening.

Thus, predictions from theoretical models may not always hold and these predictions remain to be tested. The increase of reproductive output at the range edge, for example, remains controversial (Hughes et al., 2003; Karlsson & Johansson, 2008; Bonte et al., 2012; Hudson et al., 2015). However, the allocation of resources to dispersal has been relatively well documented. For example, the fast dispersal rate and associated phenotypic traits that are observed in vanguard populations of the invasive cane toad Rhinella marina in Australia Chapter IV

(Brown et al., 2007; Alford et al., 2009; Phillips et al., 2008) provide a nice illustration of a dispersal phenotype in a rapidly expanding population.

The present study focuses on another highly and globally invasive amphibian, the African clawed frog Xenopus laevis Daudin 1802. The use of X. laevis as a model system in developmental and cellular biology (Gurdon & Hopwood, 2000) has resulted in the presence of this species in laboratories world-wide. Invasive populations of X. laevis have since become established globally as a result of accidental as well as voluntary releases from research facilities and through the release of animals from the pet trade (Measey et al., 2012).

Despite a growing body of literature on the invasion range and the impacts of this species on autochthonous ecosystems (Lafferty & Page, 1997; Lillo et al., 2005, 2011; Lobos & Jaksic,

2005; Eggert & Fouquet, 2006; Fouquet & Measey, 2006; Robert et al., 2007; Faraone et al.,

2008; Rebelo et al., 2010; Measey et al., 2012; De Busschere et al., 2016 [Appendix I]), this species has never been used to test the predictions of dispersal models. Our study focuses on an invasive population of X. laevis in the west of France. Its introduction has been suggested to be associated with the presence of a research laboratory where Xenopus were bred and maintained until the facility closed in the early 1980s (Fouquet & Measey, 2006). Animals were officially first reported in 1998 when they were observed in few ponds around the likely site of introduction. However, residents of the area subsequently suggested that animals had been in these ponds since the early 1980s. Since then, animals have been expanding at a steady rate and they now occupy an area of over 2000 square kilometers. The aim of the present study was to test whether X. laevis at the range edge show evidence of dispersal phenotypes. Specifically, we test whether animals at the range edge have longer limbs and greater locomotor performance than animals near the likely site of release. To investigate this, we analyzed terrestrial endurance capacity and limb morphology for individuals from the center and the periphery of the range.

Materials and methods

All individuals (N=164; 84 from the periphery and 80 from the center of the range) were caught in ponds and bodies of standing water within their current range using fykes. The range of X. laevis in western France is identified through regular monthly trapping campaigns by local fish and wildlife officers and currently covers three departments (Vienne, Deux-

Sèvres and Maine-et-Loire) and an area of 2000 km² (Fig.4.1). Two pairs of sites were used in this study: one site at the center of the range, near the introduction point, and one site at the periphery. For each site, all individuals were caught in a single pond. Individuals from the first pair (N=87; 53 from the periphery and 34 from the center; sites 1 and 2, respectively, in

Fig.4.1) were caught, brought back and housed at the Function and Evolution (FUNEVOL) laboratory at the Muséum National d‟Histoire Naturelle in Paris, France. Specimens were housed in groups of 6–10 individuals in 50 liter aquaria at a temperature of 23°C and fed with beef heart and mosquito larvae. All individuals were pit-tagged (Nonatec, Lutonic

International, Rodange, Luxembourg), allowing unambiguous identification of each individual during study. Individuals from the second pair of sites (N=77; 31 from the periphery and 46 from the center; sites 3 and 4, respectively, in Fig.4.1) were killed in the field using an overdose of anesthetic (MS222) according to institutional guidelines, preserved in formaldehyde and used for morphometric analyses.

Chapter IV

Fig.4.1 Current distribution of the invasive population of Xenopus laevis in the west of France.

Indicated are the point of introduction and the four sites used in this study. Small dots indicate ponds where X. laevis is present.

Tab.4.1 Morphometric traits (means±s.e.) measured for each site and sex. Site 1, N=14 males and

N=20 females; site 2, N=33 males and N=20 females; site 3, N=31 males and N=15 females; site 4,

N=20 males and N=20 females.

Sex Site 1 (core) Site 2 (edge) Site 3 (core) Site 4 (edge) SVL (mm) Female 83.37±1.03 65.01±1.03 86.70±1.04 80.54±1.04 Male 70.15±1.04 59.84±1.02 69.50±1.03 63.53±1.04 Femur length (mm) Female 27.93±1.02 28.84±1.01 27.73±1.02 27.80±1.01 Male 29.65±1.02 30.34±1.01 30.41±1.02 31.48±1.02 Tibia length (mm) Female 24.32±1.03 26.24±1.02 24.38±1.08 25.23±1.05 Male 27.35±1.03 26.61±1.02 25.59±1.07 23.71±1.07 Astragalus length (mm) Female 15.67±1.03 17.82±1.02 14.55±1.03 15.45±1.02 Male 18.54±1.03 18.28±1.02 15.14±1.03 15.31±1.03 Longest toe length (mm) Female 26.73±1.02 26.67±1.02 30.06±1.26 31.55±1.01 Male 29.11±1.02 28.12±1.02 31.05±1.02 32.66±1.02 Humerus length (mm) Female 10.59±1.04 11.86±1.03 11.22±1.32 12.02±1.19 Male 13.49±1.04 13.21±1.03 20.51±1.29 13.15±1.29 Radius length (mm) Female 9.89±1.04 9.68±1.03 12.05±1.03 13.58±1.02 Male 12.13±1.04 11.80±1.03 11.80±1.03 14.39±1.03 Hand length (mm) Female 4.00±1.04 4.31±1.03 4.52±1.05 5.16±1.03 Male 4.41±1.03 4.38±1.03 4.41±1.05 5.19±1.05 Longest finger length (mm) Female 10.02±1.03 9.35±1.02 11.25±1.03 12.42±1.02 Male 11.04±1.03 10.35±1.02 10.84±1.03 12.94±1.03 Ilium width length (mm) Female 15.78±1.02 16.63±1.02 15.07±1.03 15.10±1.02 Male 15.17±1.02 16.07±1.02 15.31±1.02 15.59±1.03 Mass (g) Female 29.51±1.03 36.14±1.03 28.84±1.06 26.915±1.04 Male 30.13±1.03 36.64±1.02 27.54±1.06 25.76±1.06

Chapter IV

Morphometrics

All individuals were weighed (Ohaus, Brooklyn, NY, USA; precision±0.1 g) and measured using a digital caliper (Mitutyo; precision±0.01 mm). Body dimensions were measured (See

Appendix J for a complete description of the measurements) following Herrel et al. (2012). A summary of the morphometric data is provided in Tab.4.1.

Fig.4.2 Top view of the track used to perform endurance tests.

Performance

Stamina tests were performed at 22°C, which is considered the optimal temperature for the species (Casterlin & Reynolds, 1980; Miller, 1982). Animals (sites 1 and 2) were placed in individual containers with some water for 1 h in an incubator set at 22°C prior to each test.

The body temperature (Tb) of each individual was recorded using a K-type thermocouple before and after stamina trial as the room temperature was slightly lower than the temperature of the incubator (19°C) causing the animals‟ body temperature to drop slightly during the trials. Between trials, animals were returned to their aquaria, fed, and allowed to rest for at least 24 h. Three trials per individual were performed and only the single best trial was

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retained for further analysis. Stamina was measured by chasing each individual down a 3 m long circular track covered with cork (Fig.4.2). Animals were chased until exhaustion, which was identified by the lack of a righting response. Note that individuals recovered quickly from these trials and were immediately ready to eat when placed back in their home aquaria. For each individual, we recorded both the total distance covered and time spent moving until exhaustion. Statistical analyses were performed using the maximum distance covered and the maximum time spent moving for each individual out of the three trials (Tab.4.2).

Chapter IV

Tab.4.2 Mean performance trait and body temperature at the end of the trial. Center, N=14 males and N=20 females; periphery, N=33 males and N=20 females.

Population Sex Mean±SE Tb (°C) Core Female 22.28±1.01 Male 21.83±1.01 Edg e Female 20.89±1.01 Male 21.63±1.01 Distance (cm) Core Female 1241.65±1.12 Male 1137.63±1.12 Edge Female 1300.17±1.11 Male 1721.87±1.09 Time (s) Core Female 75.51±1.12 Male 91.62±1.11 Edge Female 110.15±1.10 Male 148.59±1.08

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Statistical analyses

All data were log10 transformed to meet assumptions of normality and homoscedasticity. To test for differences in size [snout–vent length (SVL)] between sexes, and between center and edge sites, univariate analyses (ANOVAs) were performed. Differences in body mass, the morphology of the ilium and limb dimensions were tested between sexes and populations using multivariate ANOVAs with the SVL as a covariate (MANCOVA). These analyses were performed independently within each pair of sites (comparison of site 1 with site 2, and site 3 with site 4) to avoid potential biases due to preservation of the animals (sites 3 and 4). An

ANOVA was performed to test for differences in the body temperature of the animals from sites 1 and 2 after the trials. Given that both SVL (sites 1 and 2: F1,83=47.92; P<0.01; sites 3 and 4: F1,73=6.35; P=0.01) and body temperature (F1,83=20,65; P<0.001) were different between animals from different sites they were incorporated as covariates in our multivariate analyses. Next, a multivariate analysis (MANCOVA), with SVL and body temperature as covariates, was performed to test whether stamina, identified here as the maximum time and the maximum distance moved until exhaustion, differed between sexes and sites. All analyses were performed using SPSS v.22 (IBMSPSS, Chicago, IL, USA).

SVL was also significantly different between populations, with individuals from the center being larger than individual from the periphery (sites 1 and 2: F1,83=47.92; P<0.01; sites 3 and

4: F1,73=6.35; P=0.01). Individuals from site 1 (center of the range) are on average 20% larger than those from site 2 (periphery) (Tab.4.1, Fig.4.3). Individuals from site 3 (center of the range) are on average 8% larger than those from site 4 (periphery; Tab.4.1, Fig.4.3). Chapter IV

Results

Snout–vent length was significantly different between males and females from sites 1 and 2

(F1,83=18.80; P<0.01), with females being on average 16% larger than males. Similarly, females were on average 20% larger when comparing preserved animals from sites 3 and 4

(F1,73=50.32; P<0.01; Tab.4.3).

Tab.4.3 Results of univariate analyses testing for differences in SVL and body temperature at the end of the stamina trial between populations and sexes.

Variable Source F d.f. Error p-value SVL Population 47.92 1 83 <0.01 Sex 18.80 1 83 <0.01 Population×Sex 2.18 1 83 0.14

Tb Population 20.65 1 83 <0.01 Sex 0.88 1 83 0.35 Population×Sex 12.51 1 83 <0.01

Size-corrected limb dimensions were also significantly different between males and females from sites 1 and 2 (Wilks‟ lambda=0.53; F11,72=5.84; P<0.01; Tab.4.4), and sites 3 and 4

(Wilks‟ lambda=0.59; F14,59=2.88; P<0.01; Tab.4.5). Within sites 1 and 2, all morphological traits except for hand length were significantly greater for males (Tab.4.4). Male forelimbs are on average 15% longer and hind limbs 7% longer than those of females of a given body size

(Tab.4.1).

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Fig.4.3 Snout–vent length for males and females from sites at the center and the periphery of the range. Values are means±s.d. Site 1, N=14 males and N=20 females; site 2, N=33 males and N=20 females; site 3, N=31 males and N=15 females; site 4, N=20 males and N=20 females.

Chapter IV

Tab.4.4 MANCOVA performed on the morphometric data from sites 1 and 2 with SVL as covariate.

Effect Variable Wilks' lambda F d.f. Error p-value Population 0.61 4.25 11 72 <0.01 Mass 40.48 1 82 <0.01 Astragalus 4.56 1 82 0.04 Ilium width 6.00 1 82 0.02 Finger 6.55 1 82 0.01 Sex 0.53 5.84 11 72 <0.01 Femur 11.37 1 82 <0.01 Tibia 7.71 1 82 0.01 Astragalus 16.75 1 82 >0.01 Longest toe hind 11.49 1 82 >0.01 Humerus 26.00 1 82 >0.01 Radius 33.09 1 82 >0.01 Longest toe front 18.56 1 82 >0.01 Hand 2.43 1 82 0.12 PopulationxSex 0.79 1.73 11 72 0.08

Tab.4.5 MANCOVA performed on the morphometric data from sites 3 and 4 with SVL as covariate.

Effect Variable Wilks' lambda F d.f. Error p-value Population 0.59 2.88 14 59 <0.01 Femur 38.68 1 72 <0.01 Sex 0.48 4.66 14 59 <0.01 Toe 4.34 1 72 0.04 Radius 23.61 1 72 <0.01 Hand 7.53 1 72 <0.01 Finger 16.03 1 72 <0.01 PopulationxSex 0.78 1.19 14 59 0.31

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For sites 3 and 4, the toe, radius, hand and finger length are significantly longer in males compared with females (Tab.4.1). Limb dimensions were significantly different between populations as well (sites 1 and 2: Wilks‟ lambda=0.61; F11,72=4.25; P<0.01; Tab.4.4; sites 3 and 4: Wilks‟ lambda=0.48; F14,59=4.62; P<0.01; Tab.4.5). Individuals from site 2 (the range edge) have significantly longer astragali (on average 5% longer), a wider ilium (on average

5% wider) and a higher body mass (on average 18% heavier) than individuals from site 1

(center). Individuals from site 4 (periphery) have significantly longer femurs (on average 10% longer; Tab.4.1 & Tab.4.5) than individuals from site 3 (center). Stamina, with body length and temperature as covariates, is significantly different between males and females (Wilks‟ lambda=0.91; F2,80=4.11; P<0.02), with males being capable of moving an average of 23% longer for a given body size and temperature (Tab.4.6). Stamina is also significantly different between individuals from the center and the periphery (Wilks‟ lambda=0.83; F2,80=8.09;

P<0.01), with individuals from the range edge moving 35% longer before exhaustion (Fig.4.4,

Tab.4.6). Moreover, the distance moved also showed a trend for animals from the periphery to move a greater distance for a given body size compared with animals from the center of the range (P=0.06; Tab.4.6).

Tab.4.6 MANCOVA performed on the performance traits with SVL and temperature as covariates.

Effect Variable Wilks' lambda F d.f. Error p-value Population 0.83 8.09 2 80 <0.01 Distance 3.74 1 81 0.06 Time 15.08 1 81 <0.01 Sex 0.91 4.11 2 80 0.02 Distance 0.90 1 81 0.35 Time 6.75 1 81 0.01 PopulationxSex 0.95 2.27 2 80 0.11

Chapter IV

Fig. 4.4 Mean time until exhaustion for males and females from populations at the center versus the periphery of the range. Values are means±s.d. Site 1, N=14 males and N=20 females; site 2,

N=33 males and N=20 females.

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Discussion

Our results show significant differences in stamina and limb morphology for two sites in the range of X. laevis in France, one from the center and one on the edge of the range. Individuals at the range edge showed greater stamina and had longer legs. Our analyses of the limb morphology in a second set of populations also show longer limbs for animals from the range edge, suggesting that this is a more general phenomenon. However, measurements of locomotor performance in frogs from additional sites are needed to better understand whether the longer limbs observed in animals from these additional sites also result in differences in endurance capacity.

In addition, our results highlight morphological and locomotor differences between males and females in X. laevis. Females are significantly larger than males (Fig.4.3, Tab.4.2) as is common in frogs and in Xenopus species more specifically (Zug, 1978; Measey & Tinsley,

1998; Herrel et al., 2012). Males also possess relatively longer forelimbs and hind limbs than females (Tab.4.2). Furthermore, males show a better endurance capacity (specifically the time until exhaustion) than females (Fig.4.4, Tab.4.6) similar to observations for Xenopus tropicalis (Herrel et al., 2012). The greater endurance capacity in males relative to females may be explained by their relatively larger limbs and lower body mass allowing them to keep moving longer than females for their body size. This could be beneficial for males during courting and reproduction. The longer forelimbs observed in males may additionally provide males an advantage during mating, allowing them to maintain their grasp on females during amplexus (Measey & Tinsley, 1997). In addition to the differences between sexes, our data also show significant differences in body size and endurance capacity between the center and the periphery of the range (Fig.4.4, Tab.4.6). Individuals from the edge of the expansion range are smaller and have a higher endurance than those from the center of the range. Our results also highlight differences in body dimensions, particularly in limb segments involved in Chapter IV locomotion, such as the astragalus and ilium for individuals from site 2, and the femur for individuals from site 4. The difference in the specific skeleton elements impacted is intriguing and may be due to specific differences between populations. Whereas individuals from site 2 got to this locality by overland migration, animals from site 4 may have been taking advantage of waterways to reach their current site. Although the limb segments involved are different, these results suggest a common response: hind limbs are longer in individuals from the periphery, which is likely to enhance their locomotor capacity. Indeed, previous studies have demonstrated a relationship between limb dimensions and locomotor performance.

Higher endurance abilities are attributed to relatively longer hind limbs in X. tropicalis

(Herrel et al., 2012) and R. marina (Phillips et al., 2006). The length of the foot could also play a key role in aquatic locomotion, because foot size and rotation are crucial for the generation of thrust in aquatic frogs (Richards, 2010). The ilium may also play an important role in aquatic locomotion, as it has been suggested that the sliding of sacral vertebrae along the ilia during swimming improves swimming speed in Xenopus frogs (Videler & Jorna,

1985). Therefore, in addition to having a better terrestrial locomotor endurance, frogs from the range edge may also have an improved swimming performance compared with individuals from the center of the range. However, this remains to be tested for individuals from the four sites included in this study. From the perspective of conservation biology, it is of particular importance to pay attention to the rapid evolution of morphological and physiological traits observed in X. laevis. As highlighted in the case of the invasion of R. marina in Australia

(Brown et al., 2007; Alford et al., 2009; Phillips et al., 2006, 2008), the fast optimization of dispersal abilities can lead to situations that can be unmanageable for conservation biologists.

Our study demonstrates an improved terrestrial locomotor performance at the range edge but also suggests better locomotor abilities in an aquatic environment. In order to optimize the conservation efforts and the preservation of autochthonous ecosystems, a better understanding

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of the physiological, evolutionary and behavioral responses of invasive species that can impact dispersal and colonization is key.

This study showed significant differences in performance and morphology in X. laevis from sites in the center versus the periphery of the range. Moreover, these differences have evolved since their introduction less than 40 years ago. This suggests that, as in other invasive amphibians, spatial sorting has resulted in the evolution of locomotor capacity, improving the dispersal ability of individuals on the range edge. Although experiments are needed to test the genetic basis of these differences, the fact that there is more than 15 km between the sites from the center to the periphery suggests that gene flow may be limited and thus these subpopulations may have diverged significantly. Finally, our results are consistent with models predicting the allocation of resources to dispersal at the range edge of expanding populations. However, it remains to be tested whether this implies trade-offs with other traits such as reproductive investment, immunity or competitive ability.

Chapter VI

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CHAPTER V

Changes in dispersal allocation at the range edge of an

expanding population.

Chapter VI

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Modifications dans l’allocation à la dispersion sur le front de colonisation d’une population en expansion.

Courant J, Secondi J, Guillemet L, Vollette E, Herrel A (under review) Changes in dispersal allocation at the range edge of an expanding population. Proceedings of the Royal Society B.

Résumé

Pour les individus appartenant à une population en expansion, l‟allocation des ressources aux traits d‟histoire de vie est supposée se modifier rapidement après la colonisation de nouvelles zones. Comprendre ces changements est d‟une importance cruciale pour prédire les futures expansions et les impacts d‟une population en expansion sur les environnements colonisés.

Les études théoriques comme les données empiriques ont reportées une augmentation de l‟allocation à la dispersion dans des populations en expansion. Cet augmentation est supposée partiellement reliée au phénomène du « spatial sorting » (tri spatial) ainsi qu‟à l‟adaptation locale. Dans la présente étude, nous utilisons une population en expansion du Xénope lisse,

Xenopus laevis pour tester si les individus occupant le front de colonisation montrent une augmentation de la dispersion. Nous utilisons deux méthodes complémentaires, des expérimentations et un suivi de terrain, pour estimer le taux et la distance de dispersion. Les deux méthodes utilisées indiquent une augmentation de la dispersion. Ce résultat confirme ceux obtenus pour les populations en expansion chez d‟autres espèces, et suggère que même une population établie relativement récemment peut montrer des signaux de changements dans l‟allocation des ressources aux traits d‟histoire de vie.

Chapter VI

Changes in dispersal allocation at the range edge of an expanding population.

Courant J, Secondi J, Guillemet L, Vollette E, Herrel A (under review) Changes in dispersal allocation at the range edge of an expanding population. Proceedings of the Royal Society B.

Abstract

For individuals belonging to expanding populations, the allocation of resources to life-history traits is expected to change rapidly after the colonization of a new area. Understanding these changes is of crucial importance to predict the future range spread and the impacts of an expanding species on the colonized environments. Both theoretical and empirical studies have reported or suggested an increased dispersal at the range edge of expanding populations. This increase is thought to be partially driven by spatial sorting and adaptation. In this study we use an expanding invasive population of the African clawed frog, Xenopus laevis, to test whether at the range edge individuals show increased dispersal. We use two complementary methods, experiments and field surveys, to estimate dispersal rate and distance. Both methods indicate an increase in dispersal. This result confirms those obtained for expanding populations of other species, and suggests that even a relatively recently established populations can exhibit changes in resource allocation to life history traits.

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Introduction

On a global scale, an increasing number of species face changes in their distributions. Range shifts can be the direct consequences of human activities, i.e. through the massive increase in global transport and trade (Channell & Lomolino, 2000; Herrel & van der Meijden, 2014), or their indirect consequences, for instance through global climate change (Parmesan & Yohe,

2003). Studying the mechanisms underlying range shifts is crucial to understand and predict the dynamics of a species facing the expansion or contraction of its range.

Ranges shifts are essentially driven by three parameters summarizing the trade-offs in the life- history traits of an individual in a population: density dependence, dispersal, and population growth (Burton et al., 2010). These traits influence the probability for an individual to reproduce, disperse, and survive. According to simulation and empirical studies, the resource allocation to these traits is likely to change during the range expansion of a population, resulting in different probabilities for an individual to reproduce, disperse, and survive depending on its location within its range, i.e. the core or the edge of the range (Burton et al.,

2010). Empirical evidence abounds concerning the increased allocation to dispersal at the range edge for many taxonomical groups (Chuang & Peterson, 2016) and a few studies reported a reduced allocation to reproduction at the range edge (e.g. Hughes et al., 2003;

Hudson et al., 2015; Chapter III). The modification of the allocation to competitive ability has also been studied (Llewellyn et al., 2011; Brown et al., 2015). As underlined in previous studies on the dispersal behavior of individuals in expanding populations, great caution is necessary not to underestimate dispersal rate at the range edge possibly leading to an under- predicted expansion rate into new areas (Lindström et al., 2013).

Chapter VI

Spatial sorting has been suggested to play a substantial role in the increased dispersal at the range edge to the detriment of local adaptation (Chuang & Peterson, 2016; Shine et al., 2011).

This hypothesis may be explored by measuring the changes in dispersal at the range edge in a small or recently expanding population, i.e. where environmental and climatic conditions do not change across the range, where the influence of local adaptation is expected to be limited.

This ideal situation does not occur for several invasive species such as the cane toad in

Australia due to the age and the wide distributional range of this population.

In this study, we used an invasive and expanding population of the African clawed frog,

Xenopus laevis, introduced in France during the late 1980s (Fouquet, 2001), to assess the changes in dispersal allocation during range expansion, in a situation where, as mentioned above, the effect of local adaptation is expected to be low. This species is largely used in research laboratories all over the world and it has been introduced on four continents (Measey et al., 2012). The species has recently been classified as one of the worst invasive amphibian in the world (Measey et al., 2016).

Empirical data on overland movements and dispersal behavior is still rare (Measey, 2016), and its distribution is likely to expand greatly in Europe over the next decades (Ihlow et al.,

2016). This suggests that improving knowledge about the dispersal components exhibited by this species, i.e. dispersal distance and dispersal rate (Guillon & Bottein, 2011), is of crucial importance to predict future expansions and for the planning of future management efforts.

Moreover, an increased capacity for endurance and a longer relative length of the hind limb at the range edge of an expanding population of X. laevis has been reported (Chapter IV). Here, we tested to which extent this trait variation is related to an actual increase in dispersal components (rate and distance) in this invasive population. We used two complementary approaches to test the effect of range expansion on both dispersal components by taking into

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account the effect of i) environmental conditions in an experimental design, and ii) the landscape characteristics in a field survey. We expected that the changes in morphology and performances at the range edge (Chapter IV) were related to an actual increase in dispersal allocation. Chapter VI

Materials & Methods

Study area

Xenopus laevis was introduced in Western France in the 80s from a breeding colony in

Bouillé-Saint-Paul (Fouquet, 2001). The species has colonized a large part of the Deux-Sèvres and Maine-et-Loire departments, and expanded through the valleys of the Thouet and the

Layon, two tributaries of the Loire river (Fig.5.1). To capture frogs, fyke nets (60 cm length x

30 cm width, 6 mm mesh diameter) were introduced in ponds, with a constant capture pressure fixed at one fyke net every 50m² of water. Individuals were removed from the fyke nets and used for the experiments, or marked and released depending on the study. The research permit for both studies was provided by the Préfet of the Deux-Sèvres department.

Main principles

Two methods were applied independently to assess the changes in dispersal allocation at the range edge of the population of X. laevis by estimating the dispersal distances (D) and the dispersal rate (R). In this population, we first performed a three year survey in two areas, at the range edge and core, respectively. At the end of this survey, we sampled the ponds surrounding the survey localities in a four kilometer area to detect the individuals that had moved out of the survey ponds. On the other hand, we performed an experimental study where we recorded the overnight movements by individuals in an enclosure. In order to improve the illustration of the results, the dispersal distance, estimated by the results of the experiments, was noted DEXPE, and was transformed to vary between 0 and 1.

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Fig.5.1 Data sampling for the field survey (detailed in a and b) and the experiments.

Chapter VI

We estimated the dispersal distances with the Capture-Mark-Recapture (CMR) survey, by correcting these distances by pond density (DCMR). The dispersal rates (i.e. the proportion of individuals that moved during the experiments or the field survey) was noted REXPE and RCMR for the experiments and the field survey, respectively. To compare the relative length of the hind limb (HindR), we used the residuals of the regression of the hind limb length on the snout-vent length in a linear mixed model with the cluster and the sex of individuals as fixed effects and the site of capture as random effect. Only the moving individuals of both studies have been used for this analysis.

Common garden experiment - Sampling device

To estimate dispersal experimentally, we constructed a track composed of two corridors of eighty meters in length, seventy centimeters width and fifty centimeters high, preventing the escape of the individuals released in the track (Fig 5.2). The substrate of the device was composed of grass. In order to prevent an effect of grass height it was maintained at ten centimeters. Plastic pools (30x20x15 cm) were filled with water and placed every ten meters with the surface of pools at ground level. The water in pools was changed before every experiment with the same pond water.

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Fig.5.2 Experimental track built to estimate dispersal components. Individuals were introduced in the track at one or another extremity one heur before sunset, and the pools were checked in the morning to provide an estimate of the distance crossed by each individual in the morning.

Common garden experiment - Data collection

A total of 451 individuals were used for this experiment. Among them, 247 (116 females and

131 males) and 204 (117 females and 87 males) individuals were captured at the range core and the range edge, respectively. Individuals were captured at the range edge and the range core of the introduced population in 10 ponds for each area (Fig.5.1) during most of the period of activity (Chapter III) from April to September. During trials, individuals were introduced at one end of the device. The measurements of dispersal movement took place at night, X. laevis being mainly active at this moment (Archard, 2013). For each trial night, a group of five males was placed in a lane and five females were introduced in the other one.

Thus, intersexual interference was avoided. From one night to the other, the sex and/or the end where individuals were introduced was changed. The distance crossed by each individual was recorded in the morning, one hour after sunrise. For every individual, we noted the Chapter VI cluster (core or edge), the snout-vent length (SVL) and the length of the hind limb (See

Appendix J for a complete description of the measurements) with a digital caliper (Mitutoyo

Absolute IP67 – precision 0.01mm). Body mass was measured with a digital balance (XL 2PP

– precision 0.01g).

Before each trial, several environmental and experimental parameters were collected. The device end (DE) and the lane where individuals were introduced (DL) were noted. Taking into account these parameters in the statistical analysis may allow the identification of a potential experimental bias. The meteorological parameters were also taken into account during data sampling. For each trial night, the cloud cover (CC) was noted as a three-category factor (“1” for a cloud-free or slightly scattered sky, “2” for a half covered sky, and “3” for a completely covered sky). Similarly, for nocturnal wind strength (NWS), we noted the absence of wind as

“1” and a strong wind with “3”. The nocturnal rain (NR) was coded from “1” (no rain) to the

“3” (heavy rain). We also collected the maximal daily ambient temperature (MDT) and the moon phase. We considered the moonlight intensity (MI) following the relative brightness of the moon as a function of the lunar days (Nowinszky et al., 1979), i.e. the number of days since the last full moon.

Common garden experiment - Statistical analyzes

The effect of the capture area (range edge or range core) was tested on the dispersal rate,

REXPE, and on the dispersal distance, DEXPE, using a linear mixed model for each of these dispersal traits. Environmental (CC, WRN, MDT and MI), biological (SVL, Body mass, Sex,

HindR) and experimental (Month, DE and DL) parameters were used as covariates for both models. We added the interactions between CC and MI, MDT and WRN, and WRN and MI as covariates. The capture site was used as a random predictor. To select the parameters significantly influencing dispersal traits, we used a backward stepwise selection. As this method is severely criticized (Whittingham et al., 2006), we followed the recommendations of

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Hegyi & Laczi (2015) by using the stepwise-reintroduction for parameter estimation (SRPE).

We first performed a stepwise regression with a linear mixed model to obtain a first model with the selected parameters. Then, we tested each unselected parameter by adding it one by one to the model with the selected parameters. For parameters having a significant effect size in the new models tested, we implemented them in the final model.

CMR survey and surrounding ponds recording - sampling design

A three year capture-mark-recapture study was conducted in the population of Xenopus laevis introduced in France from 2014 to 2016. For each pond, 8 visits, starting at the beginning of

May, occurred each year with a three day interval. The method used to capture individuals is the same as the one described in the methods section. At the first visit, each captured individual was sexed, measured for its snout-vent length (SVL) and hind limb length with a digital caliper (Mitutoyo Absolute IP67 – precision 0.01mm), weighted with an electronic balance (XL 2PP – precision 0.01g), and marked with a passive integrated transponder

(Nonatec, Rodange, Luxembourg) under the skin of their back, before being released in the pond where it was captured. During the next visits, all individuals were checked for the presence/absence of a transponder. The recaptures of already marked individuals were recorded, if the previous capture occurred during the same year. New individuals were sexed, measured and marked. During the second and third year, on the first recapture of each individual, all measures (SVL, limb length, body mass) were repeated.

CMR survey and surrounding ponds recording - study sites

Two clusters of four ponds were selected at the range edge and core, respectively. The core cluster is located at approximately two kilometers from the introduction site, close to the village of Bouillé Saint-Paul, and is surrounded by a mosaic landscape dominated by pasture lands and scattered crop fields and woods. The edge cluster is located close to the village of

Saint-Georges-sur-Layon at the northern margin of the French population in the valley of the Chapter VI

Layon river, at 25 km from the introduction site. The landscape surrounding the ponds is dominated by pasture lands and vineyards, with scattered woods and crop fields. Both edge and core selected areas are crossed by many temporary streams and non-intensively used roads.

CMR survey and surrounding ponds recording - movement sampling

In 2016, after the end of the capture-mark-recapture survey, seven visits, with a one-week interval between each visit, were performed in the ponds located in a four kilometer area surrounding the core and edge clusters. Visits occurred in June and July 2016. During each visit, all captured individuals were checked for the presence/absence of a transponder under their skin. For marked individuals, SVL and body mass were recorded, as well as the limb lengths, the date and site of recapture. The estimated distances were corrected by the density of ponds in each cluster. This density was estimated by counting the ponds in two areas of 50 km² surrounding the sampling sites at the range core and the range edge, respectively. The dispersal distance, DCMR, is the result of the division of the estimated distance by the pond density of the cluster, and is expressed in kilometers per pond.

CMR survey and surrounding ponds recording - statistical analyses

To analyze differences in dispersal movements between populations of the core and the edge of the range, we used DCMR in a linear mixed model with the cluster and the biological parameters of individuals (i.e. sex, SVL and HindR) as fixed effects. We used the capture site nested in the cluster as a random effect. We also calculated the RCMR by dividing the number of records of moving individuals by the total number of individuals captured for a given cluster and a given sex.

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CMR survey and surrounding ponds recording - software

Graphics and statistical analyses were performed in R 3.2.3 (R Core Team, 2015), using the nlme package (Pinheiro et al., 2013) to perform linear mixed models with backward stepwise selections and ggplot2 (Wickham, 2009) to build the graphs. The map and the distances traveled by free-ranging frogs were respectively built and measured using the Quantum GIS software (Team QGD, 2016).

Results

Morphology

For females, the mean SVL reached 87.23 mm (± 2.29 mm SD) and 83.89 mm (± 1.50 mm) at the range core and the range edge, respectively. For males, it reached 70.86 mm (± 0.97 mm) and 69.49 mm (± 0.88 mm) at the range core and the range edge, respectively. The cluster had a significant effect on the variation in relative hind limb length of individuals measured in both studies (Linear mixed model: F1, 27= 11.03, p< 0.005), with a significantly higher relative hind limb length at the range edge (Fig.5.3). It also differed between sexes (Linear mixed model: F2, 378= 125.71, p< 0.0001), with a significantly longer relative hind limb length for males in each cluster (Fig.5.3).

Chapter VI

Fig.5.3 Relative lengths of the hind limb (limb length/snout-vent length*100) of Xenopus laevis captured at the range edge and the range core of the population introduced in Western France for animals included in the experiments and in the field surveys. Dots and error bars represent the means and the standard errors, respectively. Filled and unfilled triangles represented males and females, respectively.

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Experimental estimates of dispersal

The dispersal rate varied significantly (Tab.5.1) between core (REXPE= 62.07 %; N= 72) and edge females (REXPE= 71.79 %; N= 84). For males, the pattern was analogous, with REXPE=

61.07 % (N= 80) and REXPE= 73.56 % (N= 64) for the males of the range core and the range edge, respectively. There was no effect of the sex on REXPE, but some environmental parameters had a significant effect (Tab.5.1).

Tab.5.1 Model output for the linear mixed models run to test the effect of the cluster on REXPE and DEXPE, taking into account environmental parameters. The morphological and environmental parameters shown in the table were selected among all the parameters using a backward stepwise selection, followed by a check of the effect size of each unselected parameter.

Fixed effects Df F p

REXPE Intercept 1, 420 232,512 <0,0001 Cluster 1, 19 8,671 0,0083 Relative hind limb length 1, 420 0,971 0,3249 Maximal daily temperature (MDT) 1, 420 15,686 0,0001 Moonlight intensity (MI) 1, 420 7,712 0,0057 Nocturnal rain (NR) 2, 420 3,523 0,0304 Cloud cover (CC) 2, 420 0,827 0,4381 Sex 1, 420 2,554 0,1107 MI x CC 2, 420 2,451 0,0874

D EXPE Intercept 1, 269 1235,91 <0,0001 Cluster 1, 18 16,031 0,0008 Relative hind limb length 1, 269 11,964 0,0006 Maximal daily temperature (MDT) 1, 269 0,0001 0,9942 Nocturnal wind Strength (NWS) 2, 269 2,913 0,0561 Moonlight intensity (MI) 1, 269 0,794 0,3738 Month 1, 18 3,743 0,0689 Nocturnal rain (NR) 2, 269 3,055 0,0487 MDT x NWS 2, 269 7,991 0,0004 NWS x MI 2, 269 1,951 0,1441 Chapter VI

According to the linear mixed model performed to test the effect of the cluster on DEXPE, the variables related to the experimental device (Device-End – DE and Device-Lane – DL), the cloud cover (CC), the nocturnal rain (NR), the sex, the snout-vent length (SVL) and the date of sampling were not selected in the backward stepwise selection (Tab.5.1). The cluster effect on DEXPE was significant (Tab.5.1), with a higher DEXPE for individuals from the range edge

(Fig.5.4A). The mean DEXPE at the range edge was 14.34 % higher for males and 8.25 % higher for females compared to the range core. Nocturnal wind strength (NWS), moonlight intensity (MI), their interaction, and the interaction of maximal daily temperature (MDT) with nocturnal wind strength (NWS) had significant effects on DEXPE (Tab.5.1).

Dispersal estimates from the field survey

For females, the dispersal rate, RCMR, was significantly different between the range core and the range edge (Chi² = 10.52, df = 1, p = 0.001), with 7.69 % (12 moves) and 15.25 % (34 moves) of the females moving at the range core and at the range edge, respectively. We did not record a significant difference in the RCMR of males (Chi² = 0.28, df = 1, p = 0.559), with rates reaching 13.36 % (31 moves) and 14.40 % (27 moves) for range-core males and range- edge males, respectively.

Females travelled between 0.15 and 1.29 km at the range core and between 0.12 and 4.01 km at the range edge. Males travelled between 0.15 and 1.52 km at the range core and between

0.20 and 3.63 km at the range edge. The linear mixed model performed on the dispersal distance, DCMR, showed that the cluster effect was also significant (Linear mixed model: F1,

97= 8.67, p< 0.05), with a significantly higher DCMR at the range edge (Fig.5.4B). However, there was no effect of sex (Linear mixed model: F1, 97= 0.45, p= 0.64), nor HindR (Linear mixed model: F1, 97= 0.02, p= 0.90) on DCMR.

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Fig.5.4 Estimations of dispersal distances for the range core and the range edge of the invasive populations of Xenopus laevis in Western France. Distances were transformed to vary between 0 and 1 and are reported as DEXPE, in the experimental design (A). For the field survey, they were corrected by the density of the ponds, and are reported as DCMR (B). Dots and error bars represent the means and the standard errors, respectively. Filled and unfilled triangles represented males and females, respectively.

Chapter VI

Discussion

In this study, we report significant differences in the dispersal patterns for individuals captured at the range edge versus the range core using two different approaches. These results are based on both estimations of one-time movements from the experiments and movements based on a field survey. In the experimental study, we recorded a significant increase in both dispersal rate and dispersal distance, at the range edge. As individuals were not actively encouraged to move, this increased movement range is unlikely related to morphology alone, but also implies changes in behaviour. This partially behaviourally driven increased allocation to dispersal has also been documented for the cane toad (Gruber et al., 2017). Regardless of the species used, identifying the role of behaviour in dispersal would merit additional experimental and genetic studies to be completely understood.

For both sexes, the DCMR for range-edge moving individuals reached almost twice that of rang-core individuals. This result obtained from free-ranging animals support the results obtained from our experimental device and those obtained using performance tests for individuals from the same population (Chapter IV). It also confirms that the well documented phenomenon of increased dispersal at the range edge of the expanding populations in Rhinella marina (e.g. Phillips et al., 2008; Shine, 2011; González-Bernal et al., 2014; Gruber et al.,

2017) is not a phenomenon restricted to a single species, but may refer to a more general evolutionary process taking place during range expansions. The French invasive population of

X. laevis occupies only one climatic area (Joly et al., 2010), with no substantial variation in the landscape which is dominated by pastureland, vineyards and crop fields. Thus, this population is unlikely to face different environmental constraints at its range edge and core.

The changes in dispersal allocation at the range edge, reported in this study, are consequently unlikely related to natural selection, and may thus be the result of spatial sorting, as suggested in previous studies (Shine et al., 2011; Chuang & Peterson, 2016). Taking into account the

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role of this evolutionary process in the prediction of future or ongoing range expansions would be extremely important for the management of invasive species as well as species experiencing natural expansion of their range because of climate change.

The results for REXPE and RCMR do not completely corroborate each other, with an increased dispersal rate at the range edge for both sexes in the experimental device and only for females in the field survey. Independent of the capture area, the dispersal rates reported in this study are high enough to suggest that this species does not exhibit strong site fidelity, an extremely variable trait among amphibians (Gamble et al., 2007). Our results also confirm the importance of overland movements for this species. The importance of these movements was already suggested, but only a few empirical data supported this idea (Measey, 2016). This species, known as predominantly aquatic, seems finally to use overland movements within a high frequency.

According to the field survey, we suggest that females are likely to move less frequently than males at the range core, but not at the range edge. Male-biased dispersal rate has been reported for various species, and is mostly associated with sexual dimorphism, mating system, parental care or territoriality (Trochet et al., 2016). As far as we know, X. laevis does not use parental care, does not defend territories, and is not likely to modify its mating system.

Moreover, we did not record strong changes in sexual dimorphism that could explain the disappearance of the male-biased dispersal rate at the range edge. Even if X. laevis is not known as a territorial species, the possibly enhanced inter-specific competition for resources occurring at the range edge (Brown et al., 2013) might lead to an increased proportion of dispersing individuals of both sexes at the range edge. The competition for resources is indeed known to influence resource allocation to dispersal, especially in expanding populations

(Burton et al., 2010), but future empirical studies investigating the changes in competitive Chapter VI ability at the range edge are required to improve our knowledge about the role of competition in changes in dispersal allocation.

Our results, combined with previous studies carried out on dispersal (Chapter IV) and reproductive allocation (Chapter III) in the French invasive population of X. laevis, suggest that modifications in trade-offs between life history traits occur during range expansion. The high dispersal rates we report confirm the important potential for colonization recorded in

Europe for this species (Ihlow et al., 2016), and might increase during the next decade, as observed in other invasive species (Shine et al., 2011). This phenomenon is likely to favor an always accelerating expansion rate of the invasive populations of X. laevis. Beyond the case of this species, our results confirm that the enhanced dispersal allocation at the range edge of expanding populations also occurs for small and recent populations, reinforcing the idea of spatial sorting ruling the changes in trade-offs between the life-history traits.

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Chapter VI

137

CHAPTER VI

Dynamic demography and time since colonization in an expanding population. Chapter VI

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Dynamique des populations et durée depuis la colonisation dans une population en expansion.

Courant J, Secondi J, Vollette E, Adil L, Jacquet D, Louppe V, Herrel A (in prep) Dynamic demography and time since colonization in an expanding population.

Résumé

Les espèces exotiques envahissantes et le changement climatique sont deux phénomènes mondiaux conduisant à des modifications d‟aires de distribution dans de nombreux taxons. Durant ces modifications géographiques, des changements sont prédits dans les compromis évolutifs entre les différents traits d‟histoire de vie. Ainsi, sur le front de colonisation, l‟allocation des ressources à ces traits, la survie, le succès reproducteur et la dispersion, est réduite pour le premier trait et augmentée pour les deux autres. Ces modifications dans les compromis a une influence sur la dynamique des populations, et sur leurs taux d‟expansion. Dans cette étude, nous analysons les conséquences d‟une telle expansion sur la dynamique des populations d‟une population en expansion où des études portant sur les changements de compromis évolutifs ont déjà été menées. La méthode appliquée pour cette étude était composée de deux suivis par capture-marquage-recapture sur deux groupes de mares situées au centre de l‟aire de distribution et près du front de colonisation d‟une population en expansion de Xenopus laevis en France. En plus de ces suivis, des données de squelettochronologie ont été collectées dans la même population afin d‟estimer la variation dans la relation âge/taille entre le centre de l‟aire de distribution et le front de colonisation. Les estimations réalisées à partir du suivi par capture-marquage-recapture ont montré que la probabilité de survie était significativement supérieure sur les marges de l‟aire de distribution, sans différence entre les sexes. Aucune tendance n‟a pu être identifiée dans les estimations de probabilité d‟émigration. Les tailles de populations et les densités qui en découlent étaient deux fois plus importantes sur le front de colonisation qu‟au centre de l‟aire de distribution. Aucun changement n‟a pu être détecté dans les relations âge/taille, ni dans les structures d‟âge des deux groupes de sites. Ces résultats indiquent que X. laevis ne suit pas entièrement les prédictions des études théoriques. Cette espèce, en montrant une dynamique des populations et un patron d‟allocation des ressources peu communs parmi les espèces en expansion, pourraient s‟avérer d‟un grand intérêt pour l‟amélioration des modèles prédictifs dans le domaine des expansions d‟aires de répartitions.

Chapter VI

Dynamic demography and time since colonization in an expanding population.

Courant J, Secondi J, Vollette E, Adil L, Jacquet D, Louppe V, Herrel A (in prep) Dynamic demography and time since colonization in an expanding population.

Abstract

Invasive species and climate change are two global phenomena driving range shifts in a broad range of taxa. During these range shifts, changes in trade-offs between life history traits are predicted. At the range edge of expanding population, resource allocation to these traits, i.e. survival, reproduction and dispersal, is predicted to be reduced for the former and enhanced for the latter two. This modification in trade-offs has an influence on the dynamics of populations, and drives the expansion rates. In this study, we analyzed the consequences of a population expansion on its dynamics. The method consisted in two capture-mark-recapture survey applied at the range core and the range edge of an expanding population of Xenopus laevis in France. In addition to this survey, skeletochronological data was collected in the same expanding population to estimate the age/size relationship variation between the range core and the range edge. The estimates, obtained through with the capture-mark-recapture data, showed that the survival probability was significantly higher at the range edge than at the range core, without any difference between males and females. For emigration probability, no tendency was detected in the estimates. Population size estimates at the range edge, and the density we derived from this estimate, was twice the value estimated at the range core. No change was detected for the age/size regressions and the age structures of the range core and range edge clusters. These results indicate that X. laevis does not entirely follow the predictions of theoretical studies. This species, by exhibiting a less common pattern of resource allocation and population dynamics, may be of interest for the improvement of predictive models of range expansions.

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Introduction

The current biodiversity decline is driven by several threats, all connected to human activities

(Butchart et al., 2010). Besides their negative effects on biodiversity, two of these threats, global climate change and the introduction of invasive species, are also commonly accompanied by population ranges expansions where new environmental conditions are encountered (Channell & Lomolino, 2000). During the last decades, a significant attention has been paid on the mechanisms driving these expansions (reviewed in Chuang & Peterson,

2016). Such knowledge is of crucial importance to help landscape manager limiting the expansion of alien species, enhancing the expansion of reintroduced populations or estimating the impact of climate change (Mack et al., 2000; Thomas et al., 2004; Roman & Darling,

2007; Shine 2012).

In the context of range expansions due to the introduction of exotic species or due to range shifts driven by climate change, the trade-offs between life-history traits are predicted to change (Burton et al., 2010). These traits, i.e. survival, fecundity and age at maturity (Stearns,

1989), have commonly been studied in ecology and conservation biology. Dispersal has been suggested as an independent life-history trait influencing population demography (Neubert &

Caswell, 2000; Bonte & Dahirel, 2016). The trade-offs between these traits drive the dynamics of populations by influencing the probability of each individual to disperse, reproduce, and survive.

According to theoretical studies (Burton et al., 2010; Phillips et al., 2010), resource allocation to dispersal is enhanced at the range edge as is allocation to reproduction, except when the population is locally subject to the Allee effect. Allocation to competitive ability is predicted to be reduced at the range edge and favored at the range core, where intraspecific competition is more important because of higher densities and, as a consequence, where access to Chapter VI resources is reduced. These predictions about resource allocation to dispersal have been corroborated for many expanding populations belonging to a broad range of taxa (e.g. Haag et al., 2005; Phillips et al., 2006; Hassall et al., 2009; Laparie et al., 2013; Myles-Gonzalez et al., 2015; Chapter IV-V). Evidence of changing allocation to reproduction has also been documented in empirical studies, with both reductions (Hughes et al., 2003; Amundsen et al.,

2012; Chapter III) and increases (Hill et al., 1999; Rebrina et al., 2015) observed at the range edge of expanding populations.

As mentioned above, these life-history traits drive the dynamics of populations. Thus, a modification of the trade-offs in resource allocation to these traits is expected to impact the population dynamics. In this study, we present a case study of a well-studied invasive species, the African clawed frog, Xenopus laevis, to provide a direct observation of the population dynamics occurring at the range core and the range edge of an expanding population.

This species has been used for decades in pregnancy testing and as a biological model organism in developmental biology (Gurdon & Hopwood, 2000; van Sittert & Measey, 2016).

After accidental and voluntary escape events, several populations have been introduced out of the natural range and on four different continents (Measey et al., 2012). This species is mainly, but not exclusively, aquatic (Measey, 2016). It mainly feed on aquatic invertebrates, but it is also capable of capturing a broad range of prey including eggs, larvae and adults of amphibians (Chapter I). A negative effect of the presence of X. laevis on the native fauna has been suggested in Sicily (Lillo et al., 2011) and France (Chapter II). Impact assessments have classified this species among the worst invasive amphibians in terms of negative (ecological and economical) effects (Kraus, 2015; Measey et al., 2016; Kumschick et al., 2017). This species was introduced in France in the village of Bouillé Saint-Paul (Deux-Sèvres, France), during the late 1980s, after accidental escape events from a breeding center. Since then, the

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distribution range of the population has rapidly expanded to reach its current size. Following their own results (Grosselet et al., 2005; Thirion et al., 2009) and the experiences of landscape managers facing other invasive species (e.g. Phillips et al., 2006), local naturalist associations recommended to focus on the colonization front during their eradication programs. This advice was followed only partly, allowing the population to expand along on its northern part involving the colonization of two rivers, the Thouet and the Layon, two catchments of the

Loire river.

For this population, we already described in the previous chapters, the reduced allocation to reproduction (Chapter III) and the enhanced endurance capacity (Chapter IV) of individuals from the range edge. This result suggested an enhanced resource allocation to dispersal, corroborated by the results obtained in Chapters IV and V. The results obtained in these chapters partially corroborate the predictions of Burton and colleagues (2010), but, because of the reduced allocation to reproduction at the range edge, they question the existence of an

Allee effet in this part of the population. The Allee effect is defined as the decline in the fitness of individuals when the population density is low, possibly resulting in population decline or extinction (Courchamp et al., 2008). The influence of an Allee effect on resource allocation might be assessed by comparing the population dynamics, i.e. the demographic parameters (survival and emigration probability, population size), between the range edge and core.

Thus, in this study, we performed population dynamics analyses focusing on the French invasive and expanding population of X. laevis using a capture-mark-recapture survey and estimates of age/size relations. We compared individuals occupying ponds in the core area of the population with individuals occupying ponds at the range edge of the same expanding population. The results obtained in terms of survival probability, temporary emigration Chapter VI probability, population density and age structure of the population were used for this comparison. Following the predictions (Burton et al., 2010), we expected the survival probability to be reduced, and emigration probability to be enhanced at the range edge. If we extrapolate our results from chapter III about reproductive allocation, and follow the hypothesis of a reduced reproductive allocation when population is subjected to Allee effect

(Burton et al., 2010), we would expect to find a lower density at the range edge than at the core. A faster growth rate is commonly reported at the range edge of expanding populations

(e.g. Bøhn et al., 2004; Lindström et al., 2013; Bartle et al., 2013). Faster growth rates play a key role for colonizers to attain the adequate size to disperse after a shorter period (Marshall

& Keough, 2003) and are facilitated by a better access to resources (Lindström et al., 2013).

As a consequence, we also expect a faster growth rate at the range edge for the expanding population of X. laevis. Finally, according to our hypothesis concerning survival probability, we expect a different age structure of the edge population, with a smaller proportion of the oldest classes of individuals at the range edge.

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Material & Methods

Study area

The study took place in Western France, where the African clawed frog, Xenopus laevis, has been introduced during the late 1980s (Fouquet, 2001), after escape events from a breeding facility close to the village of Bouillé Saint-Paul (N 47.032460; W -0.300811). The population has then colonized the valleys of the Thouet and the Layon, has spread in the northern part of the Deux-Sèvres department, and has reached the Loire river (Fig.6.1). Data sampling occurred in this population where ponds were used for two distinct studies. Two groups of four ponds were selected to apply a Capture-Mark-Recapture (CMR) survey and 18 ponds scattered across the colonized range to perform a skeletochronological (SC) study

(Fig.6.1). The core cluster of ponds for the CMR survey was located at two kilometers from the introduction site. The edge cluster was located close to the village of Saint-Georges-sur-

Layon at the northern margin of its French distribution in the valley of the Layon. This cluster was located at 25 km from the introduction site. The map and estimates of pond surface were performed using Quantum GIS software version 2.18.10 (QGD Team, 2016).

Chapter VI

Fig.6.1 Sampled areas for the Capture-Mark-Recapture survey and the skelettochronological study.

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Trapping method

Individuals were captured in the ponds with fykes (60 cm length, 30 cm width and 6 mm mesh diameter). In each pond and for each capture event, the capture effort was constant and fixed at one fyke for 50 m² of water. To avoid any death by drowning of (invasive or native) amphibians, the traps were not completely immerged in the water.

CMR sampling design

The data analyzed in this study were collected during the CMR survey between 2014 and

2016. The period of capture was fixed during the most intense period of reproduction, according to the data provided in Chapter III, i.e. between April and June. Thus, the survey, composed of eight visits, started during the first days of May, and visits were repeated with a three days interval, to perform the eighth visit at the beginning of June. During visits, each unmarked captured individual was sexed, measured for its snout-vent length (SVL) and its hind limb length with a digital caliper (Mitutoyo Absolute IP67 – precision 0.01mm), weighted with an electronic balance (XL 2PP – precision 0.01g), and marked with a pit tag

(Nonatec, Rodange, Luxembourg) under the skin of its back. After this process, the frog was released in the pond where it was captured. Each already marked individual was simply recorded if it had already been captured at least once during the same year. For each individual, the first recapture of each year implied the complete set of measurements.

Skeletochronology sampling design

Eighteen ponds were sampled all around the distribution area of X. laevis in France to capture individuals and estimate their ages using a skeletochronological method. For each individual, sex, body mass, SVL and capture location were recorded. Phalanges were cut and maintained in ethanol (70%) before euthanasia by immersion in MS222. Samples were then sent to the laboratory for the cross-section preparation, using a Microtome (Microm HM360) following the protocol of Castanet & Smirina (1990). Some minor changes were brought to this Chapter VI protocol, i.e. the use of toluidine blue for coloring instead of hematoxylin. The histological analysis was performed using a Leica DM2700M and associated software (LAS V4.6). We counted the lines of arrested growth (LAGs) in 20 sections per individual (10 per slides). The number of LAGs was then used as an estimate of the age of individuals.

Data analysis - age structure

We separated the skeletochronological data in two subsamples, the core sample and the edge sample. This split allowed a comparison of growth rate between sexes and areas. We performed a nonlinear regression of the SVL on the estimated age following the von

Bertalanffy equation (von Bertalanffy, 1938), for each sex in each area. This equation is regularly used and because of its fit with the asymptotic growth of amphibians (Cogǎlniceanu

& Miaud, 2003). We used this equation in a nonlinear regression to estimate its components, i.e. the maximal SVL (SVLmax), the growth constant (K) and the theoretical age (t0) when there is no LAG:

−퐾 × 퐴푔푒 −푡0 푆푉퐿 = 푆푉퐿푚푎푥 × 1 − 푒

The residuals of each nonlinear regression were analyzed by looking at their autocorrelation and the normal quantile-quantile plot versus the standardized residuals. The coefficients of the equations obtained for each sex and each area (core or edge) were compared using pairwise t tests.

The comparison of age structure between the range edge and core populations surveyed with the CMR survey was performed by extrapolating the nonlinear regression to the SVL measurements made during the CMR survey using the reciprocal equation of von Bertalanffy:

푆푉퐿 푙표푔 1 − 푆푉퐿 퐴푔푒 = 푚푎푥 + 푡 −퐾 0

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For this extrapolation, the marked individuals from the CMR survey were removed from the analysis when their SVL was larger than the SVLmax. Indeed, if SVL > SVLmax, then

SVL/SVLmax > 1, and 1- SVL/SVLmax < 0, making impossible to estimate the age of these individuals.

Data analysis - demographic parameters

The analysis of demographic parameters was performed with the CMR data using the multistate closed robust design method (Nichols & Coffman, 1999) in the Mark software

(White & Burnham, 1999). This method assumes that no significant movements and mortality occur between the first and the last visit of the same year in the sampled area (Nichols &

Coffman, 1999). Population size was estimated using the „appropriate‟ function of the

Capture program in Mark software, enabling, thanks to Goodness Of Fit tests, to determine which source of variation (heterogeneity between individuals, time between visits, behavior related to the group) mostly influence capture probability to estimate accurately population size. The estimates were then divided by the sum of the estimated surfaces of the ponds of each surveyed area.

The multistate closed robust design models performed for each population allowed the estimation of demographic parameters by distinguishing the intervals between the primary sessions (i.e. between years) and between the secondary sessions (i.e. between visits) and by considering the ponds of capture as states. These demographic parameters consist in the survival probability, the temporary emigration probability and the capture probability.

Separate analyses were performed for the range core survey and the range edge survey, respectively. We considered that the capture probability was constant between primary and secondary sessions, assuming that the capture method did not induce positive or negative trap responses through time. We considered that the behavior of each sex and that the characteristics of the ponds could not guarantee the same capture probability between sexes Chapter VI and ponds. We considered the survival probability as different for males and females, but constant during the survey. We considered the emigration probability as the probability to leave a given pond and reach any pond. It resulted in a unique estimate of emigration probability for each pond (A, B, C and D).

Results

The nonlinear regressions performed for the core and edge individuals of each sex led to the estimates following the von Bertalanffy equation (Tab.1). We tested for differences between the parameters of these nonlinear equations, and found no significant differences, neither between areas for males (t test: t2 = -0.667, p = 0.574) and females (t test: t2 = 0.963, p =

0.437), nor between the sexes in the core (t test: t2 = 0.925, p = 0.453) and the edge area (t test: t2 = 0.911, p = 0.458).

Tab.6.1 Parameters of the nonlinear regressions of the age of individuals by their snout-vent length. The estimated parameters are the maximal snout vent length (SVLmax), the growth rate (K) and the age at zero line of arrested growth (t0).

Sex Population SVLmax (±SE) K (±SE) t0 (±SE) Females Core 99.946±10.185 0.528±0.315 -0.800±1.017 Edge 113.709±26.207 0.340±0.280 -0.955±1.259 Males Core 74.455±2.028 1.168±0.513 -0.134±0.595 Edge 70.524±3.016 1.757±1.209 -0.258±0.491

The equations obtained for each sex and for both areas were extrapolated to estimate the age structure of individuals in the surveyed ponds of the CMR survey (Fig.6.2). This extrapolation resulted in a logarithmic function with an increase in the proportion of individuals between

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one and two years old at the range core, followed, in both areas, by a consistent decrease for each category after two years of age.

For both sexes, estimating population size was not possible in the core area in 2015 (Tab.2).

The model selection in the Capture program did not lead to the selection of an estimator because of the significant effect of time, behavior and heterogeneity in the variability of the capture history of individuals.

Fig.6.2 Age structure estimates for the core (brown) and edge (orange) populations for females

(filled symbols) and males (open symbols) according to the nonlinear regressions derived from the skeletochronological data. Chapter VI

Tab.6.2 Estimates of population sizes and recapture rates according to the Capture-Mark-

Recapture survey. The population size estimates are given with the standard error. The density is the ratio between the population size and the total surface of the surveyed pond in each cluster (in m²), and is expressed in individuals per m².

Sex/Population Year Population size (± SE) Recapture rate Surface Density Females Core 1004 2014 58 ± 10.158 0.114± 0.027 0.058 2015 / 0.222e-15± 0.119e-8 / 2016 62 ± 17.016 0.096± 0.032 0.062 Edge 1921 2014 2202 ± 21848.068 0.071±0.015 1.146 2015 162 ± 17.603 0.120±0.018 0.084 2016 236 ± 28.772 0.120±0.021 0.123 Males Core 1004 2014 86 ± 8.30 0.156± 0.024 0.086 2015 / 0.002± 0.002 / 2016 45 ± 3.786 0.241± 0.036 0.045 Edge 1921 2014 2156 ± 11781.998 0.044±0.011 1.122 2015 331 ± 66.027 0.097±0.015 0.172 2016 221 ± 29.684 0.150±0.025 0.115

In 2015, the recapture rate at the range core reached its minimal value, with a probability of recapturing a marked individual below 0.01 for each visit and sex (Tab.3). In 2014, the model selection process of the population size estimate for the edge area provided, for both sex, the highest estimates and confidence intervals coupled with the lowest recapture rates of the survey in this cluster.

According to the confidence intervals, survival probability did not vary between females and males from the same area (Fig.6.3A). At the opposite, survival varied from an area to another, with lowers values for the core compared to the edge area. No clear tendency was observed for the emigration probabilities obtained during the analyses (Fig.6.3B).

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Tab.6.3 Capture probabilities parameters estimated using the multistate open robust design method. The selected model estimated capture probabilities for each sex and each pond separately.

Capture probabilities are available in Table S1.

Core Edge Females Males Females Males Capture probability pond A 0.283±0.076 0.299±0.047 0.187±0.030 0.231±0.02 9 pond B 0.159±0.068 0.063±0.027 0.265±0.047 0.046±0.048 pond C 0.051±0.032 0.133±0.035 0.282±0.061 0.197±0.057 pond D 0.127±0.051 0.148±0.043 0.051±0.033 <0.001

Chapter VI

Fig.6.3 Demographic parameters, i.e. the survival probability (A) and the emigration probability

(B), estimated using a multistate closed robust design method for males (open triangles) and females (filled triangles) from the core (brown) and the edge (orange) populations. The selected model estimated survival probability for each sex and for all surveyed sites without distinction between ponds, and estimated emigration probability for each sex and each pond separately.

Emigration probabilities are ordered for each sex and area.

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Discussion

The results of our study suggest a large difference between survival probabilities at the range core and edge, but not in the way suggested by theoretical predictions (Burton et al., 2010).

We indeed detected a higher survival probability at the range edge, suggesting a lower degree of competition, or a better competitive ability in this area, whereas empirical evidence published for expanding populations of other species reported a lower competitive ability

(Llewellyn et al., 2012; Kolbe et al., 2014; Therry et al., 2014). Thus, this result is interesting, and may find its explanation in the other results from our study. If we exclude the estimates of population size of 2014 for the edge area, when the recapture rate was low and resulted in an inaccurate estimate, the population density at the range edge was twice as large as the density in the range core. This difference suggests a decrease in available resources at the range core, but further studies remain necessary to assess the validity of this hypothesis. In studies about life history trait resource allocation in population experiencing range expansions, the range edge areas usually exhibit lower densities because they are colonized by only a few individuals with good dispersing abilities (Phillips et al., 2010; Shine et al., 2011). This results in the weak selection for individuals exhibiting a high competitive ability, and a high reproductive allocation, if the density is not low enough to locally induce an Allee effect

(Burton et al., 2010). In our study, density was higher at the range edge, and the emigration probability estimates of our model did not suggest a higher dispersal allocation at the range edge.

This does not corroborate our previous results reported in Chapters IV and V, nor the literature on dispersal allocation (e.g. Hill et al., 1999; Phillips et al., 2006; Bartoń et al.,

2012; Hudson et al., 2016). In previous chapters, we identified significant differences in traits associated with dispersal, with a significantly higher locomotor endurance at the range edge

(Chapter IV). We also detected movements over longer distances in experiments and in a field Chapter VI survey (Chapter V), and a hind limb morphology promoting these movements (Chapter IV-

V). All these results corroborated the idea of an enhanced dispersal allocation at the range edge. The demographic parameters were estimated with the multistate closed robust design

(Nichols & Coffman, 1999). This method allows the estimate of temporary migration, whereas dispersal can also imply permanent emigration, depending on an individual‟s site- fidelity. We suggest that the absence of difference in the temporary migration estimate we provide in our study is related to a very low site-fidelity of individuals. This explanation is speculative, but it appears as the most sensible if we consider the results obtained in Chapter

V, where around ten per cent of the marked individuals were recaptured in another pond than their initial capture site. These results were based on the capture-mark-recapture we used in this study. In Chapter V, we concluded that the dispersal distance, i.e. the distance travelled by the marked individuals, was higher at the range edge than at the range core. During our analyses, we did not take in account the known emigration of travelling individuals we reported in Chapter V, because it was technically not possible with this method on this software. Depending on their capture history, the travelling individuals were possibly considered as dead or temporary emigrating by the software, or present in the surveyed pond but not captured (imperfect capture probability). Future analyses should take into account the known fate of these individuals to improve the estimates of demographic parameters. This may also help provide an more accurate comparison of the core and edge respective population dynamics.

In chapter III, we detected a reduced investment in reproduction at the range edge. The reduced survival probability at the range core may be related to the cost of this increased allocation to reproduction at the range core. Thus, in our study, no strong trade-off seemed to occur between survival and dispersal allocation, but rather between reproductive and survival allocation. This result is interesting regarding the general pattern observed in the literature, i.e.

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an increased dispersal at the range edge inducing risky movements that reduce the probability of an individual to survive (e.g. Bartoń et al., 2012).

Despite the important differences in survival probability between the range core and the range edge, no differences in age structure or growth rate were detected in our analyses between these areas. According to previous empirical studies, we expected a faster growth rate at the range edge (Bøhn et al., 2004; Sanford et al., 2006; Carol et al., 2009; Lindström et al., 2013;

Bartle et al., 2013). This absence of a difference might be related to the actual time since colonization of the edge area, or to the mathematical impossibility of using large individuals in the extrapolation of the age-SVL regression. This regression was performed within the assumption that the individuals used for the skeletochronological study all stopped their growth once during one year, or if they stopped more, or less, all of them stopped it the same number of times. Environmental conditions can indeed influence the number of LAGs occurring during a year (Castanet & Smirina, 1990). It also seems important to notice that resorption of LAGs can also occur, resulting in a reduced number of LAGs and an underestimated age of individuals (Castanet & Smirina, 1990). We considered that the whole population was submitted to the same climatic conditions, preventing differences in periods of arrested growth. Moreover, the habitat where data sampling occurred did not dry up during the survey. This eliminates another source of difference in periods of arrested growth. Thus, our age estimates, even if providing relative estimates, remain valuable for comparison between areas.

The estimates performed for the demographic parameters in this study were based on the status of the edge cluster after a three year survey. This means, first, that the estimates are only based on two intervals between primary sessions. These intervals are likely to result in estimates influenced by the environmental and climatic particularities of each year. The role of these parameters is unlikely to be assessed retrospectively. Secondly, according to the Chapter VI population size estimates, the edge area was already occupied by many individuals when the survey started. This implies that the area was colonized at least several years before the beginning of the survey. These elements suggest that the population dynamics (population size and demographic parameters) in this area may have already changed since the colonization period, as well as the resource allocation to life history traits, and the trade-offs between them (Lindström et al., 2013). Despite the useful knowledge provided by our study, other approaches such as laboratory endurance tests (Chapter IV) or experimental studies

(Chapter V) were necessary to study dispersal allocation in this expanding population. It is seems important to notice that the complete CMR survey will be composed by at least five consecutive years. This may allow detecting a tendency in changing demographic parameters at the range edge, but also at the range core. The individuals occupying this area currently exhibit a very low survival probability (0.08±0.02 for both sexes), and a high reproductive allocation (Cf Chapter III). This combination may lead the population to a decreased size and density during the next years or decades.

We suggest that, regarding the changes in survival probability, traits associated with competitive ability should be experimentally assessed. For example, a comparative study of the metabolism of individuals from the range core with those from the range edge may help understanding this changing survival probability. A study on the stress response of individuals depending on the location (core or edge) they occupy in the expanding population may also be useful. This issue has already been addressed for other expanding populations exhibiting different patterns of changing trade-offs (Brown et al., 2015; Krause et al., 2016), and comparing the results between these patterns may also be helpful in the understanding of the eco-evolution of traits related to competitive ability during range expansions.

Even if the precise location of the colonization front, and the absence of cluster replicates, prevent firm conclusions about the strength of the effects we detected, the results obtained in

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our study remain interesting. They provide new insights in terms of changes in resource allocation during the colonization of new areas during range expansions. The African clawed frog seems, at least in the French invasive population, to exhibit, at the range edge an enhanced dispersal (according to Chapter IV and V) and survival probability and a reduced reproductive allocation.

Chapter VI

161

GENERAL DISCUSSION

General discussion

163

In this study, we used an expanding population of the invasive frog X. laevis to assess (i) its impact on the colonized ecosystems and (ii) the changes in resource allocation to life history traits during range expansion. The combination of these approaches aimed at providing a better understanding and assessment of the long term impacts of the species in the colonized areas in France.

In Chapter I, we described the diet of X. laevis in the French population as well as five other invasive populations and compared it with a native population. We concluded that, even if X. laevis is able to consume a wide variety of prey, the most positively selected prey categories belong to small prey items, i.e. zooplankton, dipteran larvae and small gastropods. The diversity of prey consumed was also variable between invasive and native populations as well as between invasive populations. Even if all the studied introduced populations were not likely in expansion – the Welsh population appears to have gone extinct a few years ago

(Tinsley et al., 2015) – this result was interesting. It means that this species is able to maintain a positive or stable population growth rate by feeding on a relatively small diversity of prey.

The presence of native amphibians, or other aquatic vertebrates, in the diet of several populations, suggests that X. laevis may impact amphibians by predating them. Even if studies focusing on this mechanism remain necessary, the impact of X. laevis on native amphibians has been previously suggested in Sicily (Lillo et al., 2011) and in France (Chapter II). Our study on the impact of X. laevis on native amphibians enabled us to conclude that X. laevis influenced the species richness of native amphibians by reducing the number of species in the ponds of the area it has occupied for the longest periods and where it was the most abundant.

However, our data did not enable us to conclude definitely that the invasive frog was the main or only factor influencing the decrease in species richness. Parameters related to the landscape characteristics, such as the pond density, the landscape heterogeneity, the presence of highly frequented roads or the use of pollutants disturbing amphibian reproduction may also play a General discussion key role in the species richness. The habitat preferences of X. laevis might also influence the results of the modelling we performed in Chapter II. The African clawed frog might occur at high density in ponds with characteristics that the other amphibians neglect. Determining to which extent the habitat preferences of X. laevis influences its impact on amphibian community was not possible with the dataset used in Chapter II because of the too low heterogeneity in the ponds we selected. A further study performed on the data collected during the thesis may consist in determining the habitat preferences of X. laevis. Our field observations suggest that the highest densities are found in ponds used as the water supply for cattle. The trampling of these animals prevents the development of aquatic vegetation, increases water turbidity, and eventually makes these habitats unsuitable to many aquatic invertebrates and amphibians. The African clawed frog was, however, found in large quantities in this type of ponds, suggesting that a quantitative approach may lead to a better understanding of the habitat preferences of this invasive frog. In this regard, using the same sampling design as the one used in Chapter II, but with several transects, may allow a increased heterogeneity of the habitat used, and a suitable definition of the habitat preferences of X. laevis in France.

One of the interesting the results reported in Chapter II concerns the importance of the time since colonization on the amphibian species richness. Applying this question to all the potential prey classes of X. laevis might allow an assessment of the most impacted group of prey. In this regard, recording aquatic invertebrates in ponds where we performed the study of the impact of X. laevis on native amphibians would have been of interest to understand the impact of this species on a broader range of taxa. Furthermore, our study does not focus on the ecology of native amphibians. Understanding the reasons of a potential negative effect of an invasive species necessitates a sufficient understanding of the ecology of the native species and ecosystems, which is not necessarily the case in invasion biology studies, including in

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ours. In this regard, understanding the interaction between native and invasive amphibians by determining the effect of the presence, or the abundance, of X. laevis on the body condition of adults, and on the larval development of each native co-occurring amphibian species would be helpful to understand its effects on the native amphibian community in terms of interspecific competition.

If we take into account the existing literature on the potential impact of X. laevis, this species is likely one of the most important invasive amphibians in terms of ecological impact

(Kumschick et al., 2017), and its expansion may be favoured on several continents during the next decades with climate change (Ihlow et al., 2016). Our results concerning the diet of the species and its impacts in the French invasive population strengthen the necessity of considering this invasive species as a local factor of biodiversity decline. The colonized range of X. laevis in France, and elsewhere, is subject to several other biodiversity threats. Those threats are globally much more harmful than invasive species (Millenium Ecosystem

Assessment, 2005; WWF, 2014). Thus, assessing the local relative importance of each threat on the decline of biodiversity is essential to improve the chances of success in the management of conservation programs (Brook et al., 2007).

Our results in Chapter I showed very different diets depending on the population. In a previous study, Measey (1998) demonstrated that the diet also changed according to resource availability. In Chapter II, we showed that X. laevis possibly impacted native amphibians, and that these amphibians were not necessarily the same, as those impacted by the same invasive species in Sicily (Lillo et al., 2011). These differences of ecological effects between invasive populations of the same species highlight the necessity of taking into account, to predict the long term impacts of an invasive population, and the specific interactions of a given species with the colonized environments. It strengthens the necessity of local empirical studies to General discussion perform global scale impact assessments of an invasive species, and, more generally, to assess the importance of invasive species as a major threat in the current biodiversity decline.

In the second axis of the thesis, we studied the changes in resource allocation to life history trait during the expansion of the French population of X. laevis. We followed the predictions provided in Burton and colleagues (2010), yet the results we obtained slightly differed from these predictions. Our results suggest that allocation to dispersal is enhanced at the range edge, corroborating the predictions and many empirical data published for a various range of expanding populations (e.g.Thomas et al., 2001; Simmons & Thomas, 2004; Phillips et al.,

2006; Monty & Mahy, 2010; Huang et al., 2015). We found different types of evidence that dispersal is enhanced at the range edge of this population. First, in all samples where it was measured, the morphology of the hind limb was different at the range edge compared to the range core, with a longer relative length at the range edge. This parameter has been reported as influencing the terrestrial locomotor capacities in anurans, especially in other predominantly aquatic Pipids (Herrel et al., 2014). In association to this trait, we also reported higher endurance performance at the range edge in Chapter IV, as well as a higher frequency of movements associated with longer travelled distances at the range edge in the experiments of Chapter V. In this chapter, we also reported longer movements in the wild, associated with a modified morphology (longer hind limbs) at the range edge. However, in Chapter VI, we failed to report a modified emigration probability at the pond scale between the range core and the range edge. This failure may be related to a too weak capture effort, or to a methodological weakness. In the statistical method we used, the emigration probability we estimated only concerned the temporary emigration. Information about the individuals that moved to ponds out of the surveyed sites (results shown in Chapter V) was not taken into account in this analysis. The estimates of temporary emigration might be different if the fate of individuals that moved out the surveyed area had been considered in the modeling.

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In Chapter III, we reported a decrease in reproductive allocation at the range edge. This result was predicted in theoretical studies (Burton et al., 2010) only in the case of a population subject to the Allee effect. None of our results suggest that the French population is subject to this effect, neither at the range core nor at the range edge. In Chapter III, we only took into account the ponds of a similar size and where at least thirty individuals were captured. We primarily required this sampling constraint for statistical and comparative reasons, but it also allowed us to consider that density in the edge ponds was not significantly lower at the range edge, and that individuals from the edge were unlikely to be submitted to a significantly stronger Allee effect compared to individuals from the other ponds. Such a decrease in reproductive allocation has already been reported for other expanding populations (e.g.

Hughes et al., 2003; Amundsen et al., 2012; Colautti & Barret, 2013); yet, most empirical studies predicted and reported an increased reproductive allocation at the range edge of expanding populations (e.g. Hill et al., 1999; Rebrina et al., 2015). In Chapter VI, we finally reported an increased survival probability at the range edge of the expanding population. This result suggests a higher competitive ability at the range edge compared to the range core.

Similarly to reproductive allocation, this result represented the opposite pattern to theoretical expectations (Burton et al., 2010; Chuang & Peterson, 2016) and results from empirical studies (Llewellyn et al., 2012; Kolbe et al., 2014). Even if, in our study, competitive ability was only studied through the estimate of the survival probability (which can also be influenced by other parameters, e.g. interspecific competition, climatic or meteorological events) and through the comparison of two localities, our results suggest a trade-off between life history traits. These trade-offs, at least in the expanding population of X. laevis in France, seem to follow a particular pattern: an enhanced dispersal and survival at the range edge at the expense of reproduction; and a reduced dispersal and survival at the range core, favouring reproduction (Fig.D.1). General discussion

Reproduction

Exp. Obs.

Core

Edge

Dispersal Survival

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As mentioned in Chapter VI, the survival probability is likely density-dependent. If we had estimated survival probability in sites occupied by low densities of invasive frogs, which is methodologically complicated with capture-mark-recapture surveys, our findings may have differed. As mentioned in the literature, allocation to other traits also depends on the density of conspecifics. Comparing the resource allocation to a trait, in a given area (i.e. range core or edge) of an expanding population, between sites exhibiting high and low densities would be helpful for the understanding of the role of density-dependence in resource allocation.

A possible weakness of our work relies in the absence of a comparison of resource allocation of one trait with another in the same study. As we used the same expanding population for all studies, extrapolating our general findings from each chapter to the entire population seems reasonable, but uncertainty remains concerning the quantitative aspect of the changes in trade- offs. A solution to this would consist of assessing the allocation to the life history traits for a single sample, with individuals captured at the range edge and the range core, in a sufficient number of ponds in each area. Studying the reproductive investment necessitates an immediate measurement after capture, to avoid disturbance caused by stress inducing egg laying in captivity. On the other hand, studying competitive ability – through the immune response of individuals to a stressor – implies keeping individuals alive for the experiment. As the reproductive measurement implies the euthanasia of individuals, it prevents the use of the same individuals for the reproductive investment and immune system analysis. Thus, in a first subsample, randomly selected from the total sample, individuals could be tested for their reproductive investment and their hind limb morphology. In a second subsample, individuals could be tested for their immune response to stressors and their hind limb morphology. This combination of studies would allow a direct comparison in the allocation to the life history traits, and would enable to conclude about the changing trade-offs during range expansions. General discussion

Despite these weaknesses, our results showed that the interspecific competition was reduced at the range core (Chapter II) and that the density of conspecifics in this area might be reduced as well (Chapter VI), even when no significant disturbance (captures for eradication plans) occurred during three years that could otherwise have reduced intraspecific competition. The survival probability was also reduced at the range core. These findings suggest that, at the long term, the population may be maintained at a density threshold by following a strategy where individuals allocate a consistent part of their resources to reproduction (Chapter III) instead of survival or dispersal (Chapters V-VI). The reduced survival might also be related to a decrease in resource availability at the range core. This hypothesis still needs to be assessed, but the decreased amphibian richness at the range core (Chapter II) is a first element corroborating it. Studying to which extent the diet of this species and the available resources change during the expansion of invasive populations would help solve this issue.

In every chapter of this thesis, intraspecific and interspecific competitions were not directly assessed, but appeared as crucial elements in the understanding of our results. Interspecific competition is one of the components defining the degree of threat that an invasive species represents for the other competitors, while intraspecific competition is a key factor in the dynamics of a population, whether it is invasive and expanding or not. Identifying the responses of invasive species to both types of competition in a colonized range appears as a key element in the prediction of long term impacts of these invasive species. But it also represents a true challenge, as the effects of competitions change through time and according to the characteristic of habitats.

In this thesis, we studied an expanding population for which we assumed no substantial changes in landscape (a study focusing on this issue is currently ongoing), or climate, occurred from one extremity of the range to the other (Joly et al., 2010). In this population, individuals located anywhere in the range face similar environmental conditions with

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variability mainly at the scale of the aquatic habitat. This context is not likely to favor local adaptation. Thus, as mentioned in the introduction, it is the ideal context to study the role of spatial sorting on the changes in resource allocation. One of the main particularities of spatial sorting resides in the fact that this process induces the evolution of traits in a population facing range expansion regardless of the effect of this evolution on the changes in survival probability and reproductive success (Shine et al., 2011). In our case study, the increase of dispersal at the range edge is accompanied by an increase in survival probability and a decrease of reproductive investment. Thus, the hypothesis of spatial sorting remains valid in this case study, and would deserve more replicates at the scale of the species by comparing other ongoing expanding populations, and the native population in South Africa. One of the main issues with the current knowledge of spatial sorting concerns the durability of such an evolutionary phenomenon in expanding populations. The expansion of these populations is not unlimited, and determining what happens, in terms of evolutionary processes, when the range is stabilized remains poorly known. Even if the selection of highly dispersive individuals is not driven by the survival and reproductive success of these individuals, following the spatial sorting hypothesis (Shine et al. 2011), the decrease of highly dispersive individuals after certain duration after colonization would definitely be related to their lower chances of survival and/or reproductive success. Consequently, if we consider spatial sorting as an independent evolutionary process, its importance and relevance only exist as long as other evolutionary mechanisms do not influence the modifications of life history traits. In this regard, and with our results as supporting information, studying how the evolutionary processes drive the allocation to life history traits when the influence of spatial sorting becomes insignificant is crucial to assess the evolution of traits and the long term dynamics and impacts of expanding populations. General discussion

Beyond the case of X. laevis, our results are interesting regarding the insights they bring to the issue of resource allocation during range expansion. They show that broader diversity of patterns than those currently predicted by theory are possible. Thus, predicting the long term impacts of an expanding population would first necessitate focusing on the ecology of this population, its variation through space, the eco-evolution of life history traits and the context in which it expands, instead of basing the predictions on the most common patterns observed in expanding populations.

The study of the evolutionary processes driving the changes in resource allocation to life history traits has been extensively performed in the context of expanding populations for a various range of species. In the opposite context, i.e. contracting populations, the evolutionary processes driving these changes have received far less attention despite the central importance of declining populations in conservation biology. In this context of range contractions, the nature of the threat(s) is crucial, because it influences the strength of the interspecific competition at the range margins, and the response in individual traits. Depending on the nature of the threat and on the stay or leave response of the population to a given threat, highly dispersive or highly competitive individuals could be expected at the range margins.

Understanding the evolutionary mechanisms driving the general response patterns of populations subject to a shift, in terms of life history traits, may help landscape managers by improving the predictions about the importance of the long term effects of threats and the amount of efforts needed to apply successful conservation programs.

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References

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REFERENCES

References

177

Ackerly DD (2003) Community assembly, niche conservatism, and adaptive evolution in changing environments. Int J Plant Sci 164:165–184.

Alford R a, Brown GP, Schwartzkopf L, et al (2009) Comparisons through time and space suggest rapid evolution of dispersal behavior in an invasive species. Wildl Res 36:23–28. doi:

10.1071/WR08021

Alford RA, Richards SJ (1999) Global Aamphibian declines: a problem in applied ecology.

Annu Rev Ecol Syst 30:133–165.

Amaral P, Rebelo R (2012) Diet of invasive clawed frog Xenopus laevis at Lage stream

(Oeiras, W Portugal). Herpetol J 22:187–190.

Amundsen PA, Salonen E, Niva T, et al (2012) Invader population speeds up life history during colonization. Biol Invasions 14:1501–1513. doi: 10.1007/s10530-012-0175-3

Angetter LS, Lötters S, Rödder D (2011) Climate niche shift in invasive species: The case of the brown anole. Biol J Linn Soc 104:943–954. doi: 10.1111/j.1095-8312.2011.01780.x

Baguette M, Van Dyck H (2007) Landscape connectivity and animal behavior: Functional grain as a key determinant for dispersal. Landsc Ecol 22:1117–1129. doi: 10.1007/s10980-

007-9108-4

Bálint M, Domisch S, Engelhardt CHM, et al (2011) Cryptic biodiversity loss linked to global climate change. Nat Clim Chang 1:313–318. doi: 10.1038/nclimate1191

Banci KRS, Viera NFT, Marinho PS, et al (2013) Predation of Rhinella ornata (Anura,

Bufonidae) by the alien crayfish (Crustacea, Astacidae) Procambarus clarkii (Girard, 1852) in

São Paulo, Brazil. Herpetol Notes 6:339–341. References

Barker JR, Tingley DT (1992) Air pollution effects on biodiversity. Springer Science &

Business Media, New-York

Bartle K, Moles AT, Bonser SP (2013) No evidence for rapid evolution of seed dispersal ability in range edge populations of the invasive species Senecio madagascariensis. Austral

Ecol 38:915–920. doi: 10.1111/aec.12032

Bartoń KA, Hovestadt T, Phillips BL, Travis JMJ (2012) Risky movement increases the rate of range expansion. Proc R Soc B Biol Sci 279:1194–1202. doi: 10.1098/rspb.2011.1254

Bauer JT (2012) Invasive species: “back-seat drivers” of ecosystem change? Biol Invasions

14:1295–1304. doi: 10.1007/s10530-011-0165-x

Bellard C, Bertelsmeier C, Leadley P, et al (2012) Impacts of climate change on the future of biodiversity. Ecol Lett 15:365–377. doi: 10.1111/j.1461-0248.2011.01736.x

Bellis ED (1962) The influence of humidity on wood frog activity. Am Midl Nat 68:139–148.

Berger L, Speare R, Daszak P, et al (1998) Chytridiomycosis causes amphibian mortality associated with population declines in the rain forests of Australia and Central America. Proc

Natl Acad Sci U S A 95:9031–9036. doi: 10.1073/pnas.95.15.9031

Berger L, Roberts AA, Voyles J, et al (2016) History and recent progress on chytridiomycosis in amphibians. Fungal Ecol 19:89–99. doi: 10.1016/j.funeco.2015.09.007

Bishop MJ, Peterson CH (2006) When r-selection may not predict introduced-species proliferation: predation of a nonnative oyster. Ecol Appl 16:718–730.

Bøhn T, Sandlund OT, Amundsen P, Primicerio R (2004) Changing life history during invasion. Oikos 106:138–150.

179

Bonte D, Dahirel M (2016) Dispersal: a central trait in life history. Oikos 126:472–479. doi:

10.1101/065151

Bonte D, Van Dyck H, Bullock JM, et al (2012) Costs of dispersal. Biol Rev 87:290–312. doi:

10.1111/j.1469-185X.2011.00201.x

Bosch J, Martínez-Solano I, García-París M (2001) Evidence of a chytrid fungus infection involved in the decline of the common midwife toad (Alytes obstetricans) in protected areas of central Spain. Biol Conserv 97:331–337. doi: 10.1016/S0006-3207(00)00132-4

Bosch J, Martinez-Solano I (2003) Factors influencing occupancy of breeding ponds in a montane amphibian assemblage. J Herpetol 37:410–413. doi: 10.1670/0022-

1511(2003)037[0410:FIOOBP]2.0.CO;2

Brett MT, Goldman CR (1996) A meta-analysis of the freshwater trophic cascade. Proc Natl

Acad Sci U S A 93:7723–7726.

Broennimann O, Treier UA, Müller-Schärer H, et al (2007) Evidence of climatic niche shift during biological invasion. Ecol Lett 10:701–709. doi: 10.1111/j.1461-0248.2007.01060.x

Brook BW, Sodhi NS, Bradshaw CJA (2008) Synergies among extinction drivers under global change. Trends Ecol Evol 23:453–460. doi: 10.1016/j.tree.2008.03.011

Brown GP, Shilton C, Phillips BL, Shine R (2007) Invasion, stress, and spinal arthritis in cane toads. Proc Natl Acad Sci 104:17698–17700. doi: 10.1073/pnas.0705057104

Brown GP, Kelehear C, Shilton CM, et al (2015) Stress and immunity at the invasion front: A comparison across cane toad (Rhinella marina) populations. Biol J Linn Soc 116:748–760. doi: 10.1111/bij.12623 References

Brown GP, Kelehear C, Shine R (2013) The early toad gets the worm: cane toads at an invasion front benefit from higher prey availability. J Anim Ecol 82:854–862. doi:

10.1111/1365-2656.12048

Brown GP, Phillips BL, Shine R (2014) The straight and narrow path: the evolution of straight-line dispersal at a cane toad invasion front. Proc R Soc B Biol Sci 281:1–7. doi:

10.1098/rspb.2014.1385

Buchanan BW (2006) Observed and potential effects of artificial night lighting on anuran amphibians. In: Rich C, Longcore T (eds) Ecological Consequences of Artificial Night

Lighting. Island Press, London, pp 192–220

Burton OJ, Phillips BL, Travis JMJ (2010) Trade-offs and the evolution of life-histories during range expansion. Ecol Lett 13:1210–1220. doi: 10.1111/j.1461-0248.2010.01505.x

Butchart SHM, Walpole M, Collen B, et al (2010) Global biodiversity: indicators of recent declines. Science (80- ) 328:1164–1168. doi: 10.1126/science.1187512

Cannatella DC, De Sa RO (1993) Xenopus laevis as a model organism. Syst Biol 42:426–507. doi: 10.1093/sysbio/42.4.476

Carol J, Benejam L, Benito J, García-Berthou E (2009) Growth and diet of European catfish

(Silurus glanis) in early and late invasion stages. Fundam Appl Limnol 174:317–328. doi:

10.1127/1863-9135/2009/0174-0317

Castanet J, Smirina E (1990) Introduction to the skeletochronological method in amphibians and reptiles. Ann Des Sci Nat comprenant la Zool 11:191–196. doi:

10.1017/S037689290001746X

181

Casterlin ME, Reynolds WW (1980) Diel activity and thermoregulatory behavior of a fully aquatic frog: Xenopus laevis. Hydrobiologia 75:189–191. doi: 10.1007/BF00007433

Caut S, Angulo E, Courchamp F (2008) Dietary shift of an invasive predator: rats, seabirds and sea turtles. J Appl Ecol 45:428–437. doi: 10.1111/j.1365-2664.2007.01438.x

Channell R, Lomolino M V (2000) Dynamic biogeography and conservation of endangered species. Nature 403:84–86. doi: 10.1038/47487

Chase JM, Leibold MA (2003) Ecological niches: linking classical and contemporary approaches. The University of Chicago Press, Chicago

Chuang A, Peterson CR (2016) Expanding population edges: theories, traits, and trade-offs.

Glob Chang Biol 22:494–512. doi: 10.1111/gcb.13107

Clavero M, Brotons L, Pons P, Sol D (2009) Prominent role of invasive species in avian biodiversity loss. Biol Conserv 142:2043–2049. doi: 10.1016/j.biocon.2009.03.034

Cogǎlniceanu D, Miaud C (2003) Population age structure and growth in four syntopic amphibian species inhabiting a large river floodplain. Can J Zool 81:1096–1106. doi:

10.1139/Z03-086

Colautti RI, Barrett SCH (2013) Rapid adaptation to climate facilitates range expansion of an invasive plant. Science (80- ) 342:364–366. doi: 10.1126/science.1242121

Collins JP, Storfer A (2003) Global amphibian declines : sorting the hypotheses. Divers

Distrib 9:89–98.

Colwell RK, Futuyma DJ (1971) On the measurement of niche breadth and overlap. Ecology

52:567–576. References

Comte L, Cucherousset J, Olden JD (2016) Global test of Eltonian niche conservatism of nonnative freshwater fish species between their native and introduced ranges. Ecography

(Cop) 39:1–9. doi: 10.1111/ecog.02007

Contarini M, Onufrieva KS, Thorpe KW, et al (2009) Mate-finding failure as an important cause of Allee effects along the leading edge of an invading insect population. Entomol Exp

Appl 133:307–314. doi: 10.1111/j.1570-7458.2009.00930.x

Courant J, Thirion J, Guillon M, et al (2014) Le régime alimentaire de Xenopus laevis

(Daudin, 1802) (Anura: Pipidae) introduit en France. Bull la Société Herpétologique Fr

150:1–7.

Courchamp F, Berec L, Gascoigne J (2008) Allee effects in ecology and conservation. Oxford

University Press, Cambridge.

Cree A (1989) Relationship between environmental conditions and nocturnal activity of the terrestrial frog , Leiopelma archeyi. J Herpetol 23:61–68.

Crump M (2015) Anuran reproductive modes: Evolving perspectives. J Herpetol 49:1–16.

Cushman SA (2006) Effects of habitat loss and fragmentation on amphibians : A review and prospectus. Biol Conserv 128:231–240. doi: 10.1016/j.biocon.2005.09.031

D‟Antonio CM, Dudley TL (1995) Biological invasions as agents of change on islands versus mainlands. In: Vitousek PM, Loope LL, Adsersen H (eds) Islands biological diversity and ecosystem function. Springer, Verlag, pp 103–121

183

Dalpadado P, Mowbray F (2013) Comparative analysis of feeding ecology of capelin from two shelf ecosystems, off Newfoundland and in the Barents Sea. Prog Oceanogr 114:97–105. doi: 10.1016/j.pocean.2013.05.007

Daszak P, Cunningham AA, Hyatt AD (2003) Infectious disease and amphibian population declines. Divers Distrib 9:141–150.

Davis M, Chew MK, Hobbs RJ, et al (2011) Don‟t judge species on their origins. Nature

474:153–154. doi: 10.1038/474153a

De Busschere C, Courant J, Herrel A, et al (2016) Unequal contribution of native South

African phylogeographic lineages to the invasion of the African clawed frog , Xenopus laevis, in Europe. PeerJ. doi: 10.7717/peerj.1659

Deuchar EM (1975) Xenopus: the South African clawed frog. John Wiley and Sons, London

Didham RK, Tylianakis JM, Gemmell NJ, et al (2007) Interactive effects of habitat modification and species invasion on native species decline. Trends Ecol Evol 22:489–496. doi: 10.1016/j.tree.2007.07.001

Doherty TS, Glen AS, Nimmo DG, et al (2016) Invasive predators and global biodiversity loss. Proc Natl Acad Sci 113:11261–11265. doi: 10.1073/pnas.1602480113

Duellman W, Trueb L (1986) Biology of amphibians. Johns Hopkins University Press,

Baltimore

Early R, Bradley BA, Dukes JS, et al (2016) Global threats from invasive alien species in the twenty-first century and national response capacities. Nat Commun 7:12485. doi:

10.1038/ncomms12485 References

Edmonds C, Lillie AS, Cavalli-Sforza LL (2004) Mutations arising in the wave front of an expanding population. Proc Natl Acad Sci U S A 101:975–979. doi:

10.1073/pnas.0308064100

Eggert C. & Fouquet A. (2006) A preliminary biotelemetric study of a feral invasive Xenopus laevis population in France. Alytes 23:144–149.

Essl F, Dullinger S, Rabitsch W, et al (2015) Delayed biodiversity change: No time to waste.

Trends Ecol Evol 30:375–378. doi: 10.1016/j.tree.2015.05.002

Excoffier L (2004) Patterns of DNA sequence diversity and genetic structure after a range expansion: lessons from the infinite-island model. Mol Ecol 13:853–864. doi: 10.1046/j.1365-

294X.2003.02004.x

Excoffier L, Foll M, Petit RJ (2009) Genetic consequences of range expansions. Annu Rev

Ecol Evol Syst 40:481–501. doi: 10.1146/annurev.ecolsys.39.110707.173414

Excoffier L, Ray N (2008) Surfing during population expansions promotes genetic revolutions and structuration. Trends Ecol Evol 23:347–351. doi: 10.1016/j.tree.2008.04.004

Fahrig L (2010) Effects of habitat fragmentation on biodiversity. Rev Lit Arts Am 34:487–

515. doi: 10.1146/132419

Faraone FP, Lillo F, Giacalone G, Lo Valvo M (2008) Dieta di una popolazione di xenopo liscio introdotta in Sicilia. Herpetol Sardiniae 241–245. doi: 10.13140/2.1.1500.8643

Faraone FP, Lillo F, Giacalone G, Lo Valvo M (2008) The large invasive population of

Xenopus laevis in Sicily, Italy. Amphibia-Reptilia 29:405–412. doi:

10.1163/156853808785112075

185

Ferrière R, Le Galliard JF (2001) Invasion fitness and the adaptive dynamics of space related traits. In: Clobert J, Danchin E, Dhondt A, Nichols J (eds) Dispersal. Oxford University Press,

Oxford, pp 57–79

Fisher MC, Garner TWJ (2007) The relationship between the emergence of Batrachochytrium dendrobatidis, the international trade in amphibians and introduced amphibian species.

Fungal Biol. Rev. 21:2–9.

Fisher RA (1937) The wave of advance of advantageous genes. Ann Hum Genet 7:355–369. doi: 10.1111/j.1469-1809.1937.tb02153.x

Fouquet A (2001) Des clandestins aquatiques. Zamenis 6:10–11.

Freda J (1987) The influence of acidic pond water on amphibians: a review. In: Martin HC

(ed) Acidic Precipitation: Proceedings of the International Symposium on Acidic Precipitation

Muskoka, Ontario, September 15--20, 1985. Springer Netherlands, Dordrecht, pp 439–450

Fronhofer EA, Altermatt F (2015) Eco-evolutionary feedbacks during experimental range expansions. Nat Commun 6:6844. doi: 10.1038/ncomms7844

Gales RP (1982) Age- and sex-related differences in diet selection by Rattus rattus on Stewart

Island, New Zealand. New Zeal J Zool 9:463–466. doi: 10.1080/03014223.1982.10423878

Galiana N, Lurgi M, Montoya JM, López BC (2014) Invasions cause biodiversity loss and community simplification in vertebrate food webs. Oikos 123:721–728. doi: 10.1111/j.1600-

0706.2013.00859.x References

Gamble LR, McGarigal K, Compton BW (2007) Fidelity and dispersal in the pond-breeding amphibian, Ambystoma opacum: Implications for spatio-temporal population dynamics and conservation. Biol Conserv 139:247–257. doi: 10.1016/j.biocon.2007.07.001

Gimeno I, Montserrat V, Hulme PE (2006) Are islands more susceptible to plant invasion than continents? A test using Oxalis pes-caprae L. in the western Mediterranean. J Biogeogr

33:1559–1565. doi: 10.1111/j.1365-2699.2006.01525.x

Girish S, Saidapur SK (2000) Interrelationship between food availability, fat body, and ovarian cycles in the frog, Rana tigrina, with a discussion on the role of fat body in the anuran reproduction. J Exp Zool 286:487–493.

Gkenas C, Magalhães MF, Cucherousset J, et al (2016) Long term patterns in the late summer trophic niche of the invasive pumpkinseed sunfish Lepomis gibbosus. Knowl Manag Aquat

Ecosyst 417:0–5. doi: 10.1051/kmae/2016006

Göçmen B, Cicek K, Yildiz MZ, et al (2011) A preliminary study on the feeding biology of the dice snake, Natrix tessellata, in Turkey. Mertensiella 18:365–369.

Goldberg DE (1990) Components of resource competition in plant communities. In: Grace J,

Tilman D (eds) Perspectives on Plant Competition. Academic Press Inc, San Diego, pp 27–49

González-Bernal E, Brown GP, Shine R (2014) Invasive cane toads: Social facilitation depends upon an individual‟s personality. PLoS One. doi: 10.1371/journal.pone.0102880

Grafe TU, Schmuck R, Linsenmair KE (1992) Reproductive energetics of the African reed frogs, Hyperolius viridiflavus and Hyperolius marmoratus. Physiol Zool 65:153–171.

187

Greenlees MJ, Brown GP, Webb JK, et al (2007) Do invasive cane toads (Chaunus marinus) compete with Australian frogs (Cyclorana australis)? Austral Ecol 32:900–907. doi:

10.1111/j.1442-9993.2007.01778.x

Griffith DM, Veech JA, Marsh CJ (2016) cooccur: Probabilistic species co-occurrence analysis in R. J Stat Softw 69:1–17. doi: 10.18637/jss.v069.c02

Grosselet O, Thirion J-M, Grillet P, Fouquet A (2005) Etude sur les invasions biologiques : cas du Xénope commun ou Xénope du Cap, Xenopus laevis (Daudin, 1802).

Gruber J, Brown G, Whiting MJ, Shine R (2017) Geographic divergence in dispersal-related behaviour in cane toads from range-front versus range-core populations in Australia. Behav

Ecol Sociobiol 71:1–7. doi: 10.1007/s00265-017-2266-8

Guillon JM, Bottein J (2011) A spatially explicit model of sex ratio evolution in response to sex-biased dispersal. Theor Popul Biol 80:141–149. doi: 10.1016/j.tpb.2011.04.003

Guisan A, Tingley R, Baumgartner JB, et al (2013) Predicting species distributions for conservation decisions. Ecol Lett 16:1424–1435. doi: 10.1111/ele.12189

Gulve PS (1994) Distribution and extinction patterns within a northern metapopulation of the pool frog, Rana lessonae. Ecology 75:1357–1367. doi: 10.2307/1937460

Gurdon JB, Hopwood N (2000) The introduction of Xenopus laevis into developmental biology: Of empire, pregnancy testing and ribosomal genes. Int J Dev Biol 44:43–50.

Gurevitch J, Padilla DK (2004) Are invasive species a major cause of extinctions? Trends

Ecol Evol 19:470–474. doi: 10.1016/j.tree.2004.07.005 References

Gutowsky LFG, Fox MG (2011) Occupation, body size and sex ratio of round goby

(Neogobius melanostomus) in established and newly invaded areas of an Ontario river.

Hydrobiologia 671:27–37. doi: 10.1007/s10750-011-0701-9

Haag CR, Saastamoinen M, Marden JH, Hanski I (2005) A candidate locus for variation in dispersal rate in a butterfly metapopulation. Proc R Soc B Biol Sci 272:2449–2456. doi:

10.1098/rspb.2005.3235

Hallatschek O, Nelson DR (2008) Gene surfing in expanding populations. Theor Popul Biol

73:158–170. doi: 10.1016/j.tpb.2007.08.008

Hartel T, Nemes S, Cogǎlniceanu D, et al (2007) The effect of fish and aquatic habitat complexity on amphibians. Hydrobiologia 583:173–182. doi: 10.1007/s10750-006-0490-8

Hartman G (1997) Notes on age at dispersal of beaver (Castor fiber) in an expanding population. Can J Zool 75:959–962.

Hassall C, Thompson DJ, Harvey IF (2009) Variation in morphology between core and marginal populations of three British damselflies. Aquat Insects 31:187–197. doi:

10.1080/01650420902776708

Hegyi G, Garamszegi LZ (2011) Using information theory as a substitute for stepwise regression in ecology and behavior. Behav Ecol Sociobiol 65:69–76. doi: 10.1007/s00265-

010-1036-7

Hegyi G, Laczi M (2015) Using full models, stepwise regression and model selection in ecological data sets: Monte Carlo simulations. Ann Zool Fennici 52:257–279. doi:

10.5735/086.052.0502

189

Henry RC, Bocedi G, Travis JMJ (2013) Eco-evolutionary dynamics of range shifts: Elastic margins and critical thresholds. J Theor Biol 321:1–7. doi: 10.1016/j.jtbi.2012.12.004

Herrel A, Gonwouo LN, Fokam EB, et al (2012) Intersexual differences in body shape and locomotor performance in the aquatic frog, Xenopus tropicalis. J Zool 287:311–316. doi:

10.1111/j.1469-7998.2012.00919.x

Herrel A, van der Meijden A (2014) An analysis of the live reptile and amphibian trade in the

USA compared to the global trade in endangered species. Herpetol J 24:103–110.

Herrel A, Vasilopoulou-Kampitsi M, Bonneaud C (2014) Jumping performance in the highly aquatic frog, Xenopus tropicalis: sex-specific relationships between morphology and performance. PeerJ 2:e661. doi: 10.7717/peerj.661

Hill JK, Thomas CD, Blakeley DS (1999) Evolution of flight morphology in a butterfly that has recently expanded its geographic range. Oecologia 121:165–170. doi:

10.1007/s004420050918

Hines JE (2006) Presence - Software to estimate patch occupancy and related parameters.

Hinsch M, Pen I, Komdeur J (2013) Evolution of defense against depletion of local food resources in a mechanistic foraging model. Behav Ecol 24:245–252. doi:

10.1093/beheco/ars160

Holland CA, Dumont JN (1975) Oogenesis in Xenopus laevis (Daudin). Cell Tissue Res

162:177–184. References

Huang F, Peng S, Chen B, et al (2015) Rapid evolution of dispersal-related traits during range expansion of an invasive vine Mikania micrantha. Oikos 124:1023–1030. doi:

10.1111/oik.01820

Hudson CM, Brown GP, Shine R (2016) Athletic anurans: the impact of morphology, ecology and evolution on climbing ability in invasive cane toads. Biol J Linn Soc 119:992–999.

Hudson CM, Phillips BL, Brown GP, Shine R (2015) Virgins in the vanguard: low reproductive frequency in invasion-front cane toads. Biol J Linn Soc 116:743–747. doi:

10.1111/bij.12618

Hughes CL, Hill JK, Dytham C (2003) Evolutionary trade-offs between reproduction and dispersal in populations at expanding range boundaries. Proc R Soc B Biol Sci 270:S147–

S150. doi: 10.1098/rsbl.2003.0049

Iela L, Milone M, Caliendo MF, et al (1979) Role of lipids in the physiology of the testis of

Rana esculenta: Annual changes in the lipid and protein content of the liver, fat body, testis and plasma. Bolletino di Zool 46:11–16. doi: 10.1080/11250007909440272

Ihlow F, Courant J, Secondi J, et al (2016) Impacts of climate change on the global invasion potential of the African clawed frog Xenopus laevis. PLoS One 11:e0154869. doi:

10.1371/journal.pone.0154869

Jakob EM, Marshall SD, Uetz GW (1996) Estimating fitness: a comparison of body condition indices. Oikos 77:61–67.

Joly D, Brossard T, Cardot H, et al (2010) Les types de climats en France, une construction spatiale. Cybergeo Eur J Geogr. doi: 10.4000/cybergeo.23155

191

Karlsson B, Johansson A (2008) Seasonal polyphenism and developmental trade-offs between flight ability and egg laying in a pierid butterfly. Proc Biol Sci 275:2131–6. doi:

10.1098/rspb.2008.0404

Katsanevakis S, Wallentinus I, Zenetos A, et al (2014) Impacts of invasive alien marine species on ecosystem services and biodiversity: a pan-European review. Aquat Invasions

9:391–423. doi: 10.3391/ai.2014.9.4.01

Klaver RW, Peterson CR, Patla DA (2013) Influence of water conductivity on amphibian occupancy in the greater Yellowstone ecosystem. West North Am Nat 73:184–197. doi:

10.3398/064.073.0208

Kolbe JJ, Ehrenberger JC, Moniz HA, Angilletta MJ (2014) Physiological variation among invasive populations of the brown anole (Anolis sagrei). Physiol Biochem Zool 87:92–104. doi: 10.1086/672157

Kraus F (2015) Impacts from invasive reptiles and amphibians. Annu Rev Ecol Evol Syst

46:75–97. doi: 10.1146/annurev-ecolsys-112414-054450

Krause JS, Chmura HE, Pérez JH, et al (2016) Breeding on the leading edge of a northward range expansion: differences in morphology and the stress response in the arctic Gambel‟s white-crowned sparrow. Oecologia 180:33–44. doi: 10.1007/s00442-015-3447-7

Krauss J, Bommarco R, Guardiola M, et al (2010) Habitat fragmentation causes immediate and time-delayed biodiversity loss at different trophic levels. Ecol Lett 13:597–605. doi:

10.1111/j.1461-0248.2010.01457.x

Kubisch A, Fronhofer EA, Poethke HJ, Hovestadt T (2013) Kin competition as a major driving force for invasions. Am Nat 181:700–706. doi: 10.1086/670008 References

Kumschick S, Vimercati G, de Villiers FA, et al (2017) Impact assessment with different scoring tools: How well do alien amphibian assessments match? NeoBiota 33:53–66. doi:

10.3897/neobiota.33.10376

Lachavanne J, Juge R, Perfetta J (1995) Structure des peuplements de macrophytes. In:

Pourriot R, Meybeck M (eds) Limnologie Générale. Masson, Paris, pp 473–493

Lafferty KD, Page CJ (1997) Predation on the endangered tidewater goby, Eucyclogobius newberryi, by the introduced African clawed frog, Xenopus laevis, with notes on the frog‟s parasites. Copeia 1997:589–592.

Laparie M, Renault D, Lebouvier M, Delattre T (2013) Is dispersal promoted at the invasion front? Morphological analysis of a ground beetle invading the Kerguelen Islands, Merizodus soledadinus (Coleoptera, Carabidae). Biol Invasions 15:1641–1648. doi: 10.1007/s10530-

012-0403-x

Lechowicz MJ (1982) The sampling characteristics of electivity indices. Oecologia 52:22–30. doi: 10.1007/BF00349007

Léotard G, Debout G, Dalecky A, et al (2009) Range expansion drives dispersal evolution in an equatorial three-species symbiosis. PLoS One 4:1–11. doi: 10.1371/journal.pone.0005377

Lescure J, de Massary JC (2012) Atlas des amphibiens et reptiles de France. Biotope:

Muséum national d‟histoire naturelle

Liebl AL, Martin LB (2012) Exploratory behaviour and stressor hyper-responsiveness facilitate range expansion of an introduced songbird. Proc R Soc B Biol Sci 279:4375–4381. doi: 10.1098/rspb.2012.1606

193

Lillo F, Faraone FP, Lo Valvo M (2011) Can the introduction of Xenopus laevis affect native amphibian populations? Reduction of reproductive occurrence in presence of the invasive species. Biol Invasions 13:1533–1541. doi: 10.1007/s10530-010-9911-8

Lillo F, Sicilia A, Castelli G (2005) An invasive population of Xenopus laevis (Daudin, 1802) in Italy. Herpetozoa 18:63–64.

Lindström T, Brown GP, Sisson S a, et al (2013) Rapid shifts in dispersal behavior on an expanding range edge. Proc Natl Acad Sci U S A 110:13452–13456. doi:

10.1073/pnas.1303157110

Llewellyn D, Thompson MB, Brown GP, et al (2012) Reduced investment in immune function in invasion-front populations of the cane toad (Rhinella marina) in Australia. Biol

Invasions 14:999–1008. doi: 10.1007/s10530-011-0135-3

Llewelyn J, Phillips BL, Alford RA, et al (2010) Locomotor performance in an invasive species: Cane toads from the invasion front have greater endurance, but not speed, compared to conspecifics from a long-colonised area. Oecologia 162:343–348. doi: 10.1007/s00442-

009-1471-1

Lobos G, Jaksic FM (2005) The ongoing invasion of African clawed frogs (Xenopus laevis) in

Chile: causes of concern. Biodivers Conserv 14:429–439.

Lobos G, Measey GJ (2002) Invasive populations of Xenopus laevis (Daudin) in Chile.

Herpetol J 12:163–168.

Lockwood JL (1999) Using taxonomy to predict success among introduced avifauna: relative importance of transport and establishment. Conserv Biol 13:530–567. References

Lockwood JL, Hoopes MF, Marchetti MP (2013) Invasion Ecology. John Wiley and Sons,

Oxford

Lockwood JL, Hoopes MF, Marchetti MP (2013) Invasion ecology, 2nd edn. John Wiley and

Sons, Oxford

Lodge DM (1993) Biological invasions: lessons for ecology. Trends Ecol Evol 8:133–136.

Luiselli LM, Capizzi D, Filippi E, et al (2007) Comparative diets of three populations of an aquatic snake (Natrix tessellata, Colubridae) from mediterranean streams with different hydric regimes. Copeia 2007:426–435. doi: 10.1643/0045-

8511(2007)7[426:CDOTPO]2.0.CO;2

MacCracken JG, Stebbings JL (2012) Test of a body condition index with amphibians. J

Herpetol 46:346–350. doi: 10.1670/10-292

Mack RN, Simberloff D, Lonsdale WM, et al (2000) Biotic invasions: causes, epidemiology, global consequences, and control. Ecol Appl 10:689–710.

MacKenzie DI, Nichols JD, Andrew Royle J, et al (2006) Occupancy estimation and modeling. Inferring patterns and dynamics of species occurrence. Elsevier, Oxford

Maiti S, Chowdhury A (2013) Effects of anthropogenic pollution on mangrove biodiversity: a review. J Environ Prot (Irvine, Calif) 2013:1428–1434. doi: 10.4236/jep.2013.412163

Marshall DJ, Keough MJ (2003) Effects of settler size and density on early post- settlement survival of Ciona intestinalis in the field. Mar Ecol Prog Ser 259:139–144. doi:

10.3354/meps259139

195

McCoid MJ, Fritts TH (1989) Growth and fatbody cycles in feral populations of the African clawed frog, Xenopus laevis (Pipidae), in California with comments on reproduction.

Southwest Nat 34:499–505.

McCoid MJ, Fritts TH (1980) Notes on the diet of a feral population of Xenopus laevis

(Pipidae) in California. Southwest Nat 25:272–275. doi: 10.2307/3671256

McMullin RT, Bennett LL, Bjorgan OJ, et al (2016) Relationships between air pollution, population density, and lichen biodiversity in the Niagara Escarpment World Biosphere

Reserve. Lichenol 48:593–605. doi: 10.1017/S0024282916000402

Measey GJ (1998) Diet of feral Xenopus laevis (Daudin) in South Wales, U.K. J Zool

246:287–298.

Measey GJ (2001) Growth and ageing of feral Xenopus laevis (Daudin) in South Wales, U.K.

J Zool 254:547–555.

Measey GJ (1998) Terrestrial prey capture in Xenopus laevis. Copeia 3:787–791.

Measey GJ (2016) Overland movement in African clawed frogs (Xenopus laevis): a systematic review. PeerJ 4:e2474. doi: 10.7717/peerj.2474

Measey GJ, Rödder D, Green SL, et al (2012) Ongoing invasions of the African clawed frog,

Xenopus laevis: a global review. Biol Invasions 14:2255–2270. doi: 10.1007/s10530-012-

0227-8

Measey GJ, Fouquet A (2006) Plotting the course of an African clawed frog invasion in

Western France. Anim Biol 56:95–102. References

Measey GJ, Tinsley RC (1997) Mating behavior of Xenopus wittei (Anura: Pipidae). Copeia

3:601–609.

Measey GJ, Tinsley RC (1998) Feral Xenopus laevis in South Wales. Herpetol J 8:23–27.

Measey GJ, Vimercati G, de Villiers FA, et al (2015) Frog eat frog: exploring variables influencing anurophagy. PeerJ 3:e1204. doi: 10.7717/peerj.1204

Measey GJ, Vimercati G, de Villiers FA, et al (2016) A global assessment of alien amphibian impacts in a formal framework. Divers Distrib 1–12. doi: 10.1017/CBO9781107415324.004

Millenium Ecosystem Assessment (2005) Ecosystems and human well-being: Synthesis.

Island Press, Washington, DC

Miller K (1982) Effect of temperature on sprint performance in the frog Xenopus laevis and the salamander Necturus maculosus. Copeia 1982:695–698. doi: 10.2307/1444671

Miller TEX, Inouye BD (2013) Sex and stochasticity affect range expansion of experimental invasions. Ecol Lett 16:354–361. doi: 10.1111/ele.12049

Molnar JL, Gamboa RL, Revenga C, Spalding MD (2008) Assessing the global threat of invasive species to marine biodiversity. Front Ecol Environ 6:485–492. doi: 10.1890/070064

Monty A, Mahy G (2010) Evolution of dispersal traits along an invasion route in the wind- dispersed Senecio inaequidens (Asteraceae). Oikos 119:1563–1570. doi: 10.1111/j.1600-

0706.2010.17769.x

Morisita M (1959) Measuring of interspecific association and similarity between communities. Mem Fac Sci Kyushu Univ 3:65–80.

197

Mukherjee A, Williams DA, Wheeler GS, et al (2012) Brazilian peppertree (Schinus terebinthifolius) in Florida and South America: Evidence of a possible niche shift driven by hybridization. Biol Invasions 14:1415–1430. doi: 10.1007/s10530-011-0168-7

Munir T, Hussain M, Naseem S (2016) Water pollution - A menace of freshwater biodiversity: a review. J Entomol Zool Stud 4:578–580.

Myles-Gonzalez E, Burness G, Yavno S, et al (2015) To boldly go where no goby has gone before: Boldness, dispersal tendency, and metabolism at the invasion front. Behav Ecol

26:1083–1090. doi: 10.1093/beheco/arv050

Neubert MG, Caswell H (2000) Demography and dispersal: Calculation and sensitivity analysis of invasion speed for structured populations. Ecology 81:1613–1628. doi:

10.2307/177311

Nichols JD, Coffman CJ (1999) Demographic parameter estimation for experimental landscape studies on small mammal populations. In: Barrett GW, Peles JD (eds) Landscape

Ecology of Small Mammals. Springer New York, New York, NY, pp 287–309

Nowinszky L, Szabó S, Toth G, et al (1979) The effect of the moon phases and of the intensity of polarized moonlight on the light-trap catches. Zeitschrift für Angew Entomol

88:337–353. doi: 10.1111/j.1439-0418.1979.tb02512.x

Olson DH, Aanensen DM, Ronnenberg KL, et al (2013) Mapping the global emergence of

Batrachochytrium dendrobatidis, the amphibian chytrid fungus. PLoS One 8:e56802. doi:

10.1371/journal.pone.0056802

Olson LJ (2006) The economics of terrestrial invasive species: a review of the literature.

Agric Resour Econ Rev 35:178–194. References

Pagnucco KS, Ricciardi A (2015) Disentangling the influence of abiotic variables and a non- native predator on freshwater community structure. Ecosphere 6:1–17. doi: 10.1890/ES15-

00371.1

Parmesan C, Yohe G (2003) A globally coherent fingerprint of climate change impacts across natural systems. Nature 421:37–42. doi: 10.1038/nature01286

Paunović A, Bjelić-Cabrilo O, Šimić S (2010) The diet of water frogs (Pelophylax esculentus

“complex”) from the Petrovaradinski Rit marsh (Serbia). Arch Biol Sci 62:797–806. doi:

10.2298/ABS1003797P

Peig J, Green AJ (2010) The paradigm of body condition: A critical reappraisal of current methods based on mass and length. Funct Ecol 24:1323–1332. doi: 10.1111/j.1365-

2435.2010.01751.x

Peig J, Green AJ (2009) New perspectives for estimating body condition from mass/length data: The scaled mass index as an alternative method. Oikos 118:1883–1891. doi:

10.1111/j.1600-0706.2009.17643.x

Phillips BL, Brown GP, Shine R (2010) Life-history evolution in range-shifting populations.

Ecology 91:1617–1627. doi: 10.1890/09-0910.1

Phillips BL, Brown GP, Shine R (2010) Life-history evolution in range-shifting populations.

Ecology 91:1617–1627. doi: 10.1890/09-0910.1

Phillips BL, Brown GP, Travis JMJ, Shine R (2008) Reid‟s paradox revisited: the evolution of dispersal kernels during range expansion. Am Nat 172:S34–S48. doi: 10.1086/588255

199

Phillips BL, Brown GP, Webb JK, Shine R (2006) Invasion and the evolution of speed in toads. Nature 439:803. doi: 10.1038/439803a

Pimm SL, Raven P (2000) Biodiversity. Extinction by numbers. Nature 403:843–845. doi:

10.1038/35002708

Pinheiro J, Bates D, DebRoy S, et al (2013) nlme: linear and nonlinear mixed effects models.

1–3.

Prigioni C, Balestrieri A, Remonti L (2005) Food habits of the coypu, Myocastor coypus, and its impact on aquatic vegetation in a freshwater habitat of NW Italy. Folia Zool 54:269–277.

Pyšek P, Jarošík V, Hulme PE, et al (2012) A global assessment of invasive plant impacts on resident species, communities and ecosystems: The interaction of impact measures, invading species‟ traits and environment. Glob Chang Biol 18:1725–1737. doi: 10.1111/j.1365-

2486.2011.02636.x

QGD Team (2016) Quantum GIS Geographic Information System.

R Core Team (2015) R: a language and environment for statistical computing. R Found. Stat.

Comput. Vienna, Austria 0:{ISBN} 3-900051-07-0.

Rasar MA, Hammes SR (2006) The physiology of the Xenopus laevis ovary. In: Clifton NJ

(ed) Methods in molecular biology. Humana Press Inc., Totowa, pp 17–30

Rebelo R, Amaral P, Bernardes M, et al (2010) Xenopus laevis (Daudin, 1802), a new exotic amphibian in Portugal. Biol Invasions 12:3383–3387. doi: 10.1007/s10530-010-9757-0 References

Rebrina F, Skejo J, Lucić A, Hudina S (2015) Trait variability of the signal crayfish

(Pacifastacus leniusculus) in a recently invaded region reflects potential benefits and trade- offs during dispersal. Aquat Invasions 10:41–50. doi: 10.3391/ai.2015.10.1.04

Rehage JS, Barnett BK, Sih A (2005) Foraging behaviour and invasiveness: Do invasive

Gambusia exhibit higher feeding rates and broader diets than their noninvasive relatives? Ecol

Freshw Fish 14:352–360. doi: 10.1111/j.1600-0633.2005.00109.x

Ribeiro R, Carretero MA, Sillero N, et al (2011) The pond network: can structural connectivity reflect on (amphibian) biodiversity patterns ? Landsc Ecol 26:673–682. doi:

10.1007/s10980-011-9592-4

Ricciardi A, Cohen J (2007) The invasiveness of an introduced species does not predict its impact. Biol Invasions 9:309–315. doi: 10.1007/s10530-006-9034-4

Ricciardi A, Rasmussen JB (1998) Predicting the identity and impact of future biological invaders: a priority for aquatic resource management. Can J Fish Aquat Sci 55:1759–1765. doi: 10.1890/02-0571

Richards CT (2010) Kinematics and hydrodynamics analysis of swimming anurans reveals striking inter-specific differences in the mechanism for producing thrust. J Exp Biol 213:621–

634. doi: 10.1242/jeb.032631

Robert J, Abramowitz L, Gantress J, Morales HD (2007) Xenopus laevis: a possible vector of

Ranavirus infection? J Wildl Dis 43:645–652. doi: 10.7589/0090-3558-43.4.645

Robert J, Grayfer L, Edholm ES, et al (2014) Inflammation-induced reactivation of the ranavirus frog VIRUS 3 in asymptomatic Xenopus laevis. PLoS One 9:e112904. doi:

10.1371/journal.pone.0112904

201

Rödder D, Ihlow F, Courant J, et al (2017) Global realized niche divergence in the African- clawed frog Xenopus laevis. Ecol Evol 7:4044–4058. doi: 10.1002/ece3.3010

Roman J, Darling J (2007) Paradox lost: genetic diversity and the success of aquatic invasions. Trends Ecol Evol 22:454–464. doi: 10.1016/j.tree.2007.07.002

Rothermel BB (2004) Migratory success of juveniles: a potential constraint on connectivity for pond-breeding Amphibians. Ecol Appl 14:1535–1546.

Royle JA (2004) N-mixture models for estimating population size from spatially replicated counts. Biometrics 60:108–115. doi: 10.1111/j.0006-341X.2004.00142.x

Royle JA, Nichols JD (2003) Estimating abundance from repeated presence-absence data or point counts. Ecology 84:777–790. doi: 10.1890/0012-

9658(2003)084[0777:EAFRPA]2.0.CO;2

Rutz C, Whittingham MJ, Newton I (2006) Age-dependent diet choice in an avian top predator. Proc R Soc B Biol Sci 273:579–586. doi: 10.1098/rspb.2005.3353

Saenz D, Fitzgerald LA, Baum KA, Conner RN (2006) Abiotic correlates of anuran calling phenology: the importance of rain, temperature, and season. Herpetol Monogr 20:64–82.

Sakai AK, Allendorf FW, Holt JS, et al (2001) The population biology of invasive species.

Annu Rev Ecol Syst 32:305–332.

Sanford E, Holzman SB, Haney RA, et al (2006) Larval tolerance, gene flow, and the northern geographic range limit of fiddler crabs. Ecology 87:2882–2894.

Sax D, Stachowicz J, Brown J, et al (2007) Ecological and evolutionary insights from species invasions. Trends Ecol Evol 22:465–471. doi: 10.1016/j.tree.2007.06.009 References

Schafer LN, Platell ME, Valesini FJ, Potter IC (2002) Comparisons between the influence of habitat type, season and body size on the dietary compositions of fish species in nearshore marine waters. J Exp Mar Bio Ecol 278:67–92. doi: 10.1016/S0022-0981(02)00337-4

Schoonbee HJ, Prinsloo JF, Nxiweni JG (1992) Observations on the feeding habits of larvae, juvenile and adult stages of the African clawed frog, Xenopus laevis, in impoundments in

Transkei. Water SA 18:227–236.

Schramm M (1987) Control of Xenopus laevis (Amphibia: Pipidae) in fish ponds with observations on its threat to fish fry and fingerlings. Water SA 13:53–56.

Shannon CE, Weaver W (1964) The mathematical theory of communication. University of

Illinois Press, pp 3–28

Shapiro HA, Zwarenstein H (1934) A rapid test for pregnancy on Xenopus laevis. Nature

133:762.

Shine R (2012) Invasive species as drivers of evolutionary change: cane toads in tropical

Australia. Evol Appl 5:107–116. doi: 10.1111/j.1752-4571.2011.00201.x

Shine R, Brown GP, Phillips BL (2011) An evolutionary process that assembles phenotypes through space rather than through time. Proc Natl Acad Sci U S A 108:5708–5711. doi:

10.1073/pnas.1018989108

Simmons AD, Thomas CD (2004) Changes in dispersal during species‟ range expansions. Am

Nat 164:378–395. doi: 10.1086/423430

Skellam J G (1951) Random dispersal in theoretical populations. Biometrika 38:196–218. doi:

10.1016/S0092-8240(05)80044-8

203

Slatyer RA, Hirst M, Sexton JP (2013) Niche breadth predicts geographical range size: a general ecological pattern. Ecol Lett 16:1104–1114. doi: 10.1111/ele.12140

Soberón J (2007) Grinnellian and Eltonian niches and geographic distributions of species.

Ecol Lett 10:1115–1123.

Soto-Azat C, Peñafiel-Ricaurte A, Price SJ, et al (2016) Xenopus laevis and emerging amphibian pathogens in Chile. Ecohealth 13:775–783. doi: 10.1007/s10393-016-1186-9

Stearns SC (1989) The evolution of life histories. Oxford University Press, New-York

Stevens VM, Pavoine S, Baguette M (2010) Variation within and between closely related species uncovers high intra-specific variability in dispersal. PLoS One. doi:

10.1371/journal.pone.0011123

Stiels D, Gaißer B, Schidelko K, et al (2015) Niche shift in four non-native estrildid finches and implications for species distribution models. Ibis (Lond 1859) 157:75–90. doi:

10.1111/ibi.12194

Strauss SY, Rudgers J, Lau J, Irwin R (2002) Direct and ecological costs of resistance to herbivory. Trends Ecol Evol 17:278–285. doi: 10.1016/S0169-5347(02)02483-7

Therry L, Zawal A, Bonte D, Stoks R (2014) What factors shape female phenotypes of a poleward-moving damselfly at the edge of its range? Biol J Linn Soc 112:556–568. doi:

10.1111/bij.12295

Thirion J-M, Doré F, Grillet P, et al (2009) Etude et propositions de modes de lutte contre le

Xénope lisse Xenopus laevis. References

Thomas CD, Bodsworth EJ, Wilson RJ, et al (2001) Ecological and evolutionary processes at expanding range margins. Nature 411:577–581. doi: 10.1038/35079066

Thomas CD, Cameron A, Green RE, et al (2004) Extinction risk from climate change. Nature

427:145–148.

Tillberg C V., Holway DA, Lebrun EG, Suarez A V. (2007) Trophic ecology of invasive

Argentine ants in their native and introduced ranges. Proc Natl Acad Sci U S A 104:20856–

20861. doi: 10.1073/pnas.0706903105

Tinsley RC, Coxhead PG, Stott LC, et al (2015) Chytrid fungus infections in laboratory and introduced Xenopus laevis populations: assessing the risks for U.K. native amphibians. Biol

Conserv 184:380–388. doi: 10.1016/j.biocon.2015.01.034

Tinsley RC, Stott LC, Viney ME, et al (2015) Extinction of an introduced warm-climate alien species, Xenopus laevis, by extreme weather events. Biol Invasions 17:3183–3195. doi:

10.1007/s10530-015-0944-x

Travis JMJ, Dytham C (2002) Dispersal evolution during invasions. Evol Ecol Res 4:1119–

1129.

Travis JMJ, Münkemüller T, Burton OJ (2010) Mutation surfing and the evolution of dispersal during range expansions. J Evol Biol 23:2656–2667. doi: 10.1111/j.1420-

9101.2010.02123.x

205

Trochet A, Courtois EA, Stevens VM, et al (2016) Evolution of Sex-Biased Dispersal. Q Rev

Biol 91:297–320. doi: 10.1086/688097

Van Ngo B, Lee YF, Ngo CD (2014) Variation in dietary composition of granular spiny frogs

(Quasipaa verrucospinosa) in central Vietnam. Herpetol J 24:245–253. van Sittert L, Measey GJ (2016) Historical perspectives on global exports and research of

African clawed frog (Xenopus laevis). Trans R Soc South Africa 71:157–166. doi:

10.1080/0035919X.2016.1158747

Vanderploeg HA, Scavia D (1979) Calculation and use of selectivity coefficients of feeding:

Zooplankton grazing. Ecol Modell 7:135–149. doi: 10.1016/0304-3800(79)90004-8

Veech JA (2013) A probabilistic model for analysing species co-occurrence. Glob Ecol

Biogeogr 22:252–260. doi: 10.1111/j.1466-8238.2012.00789.x

Videler JJ, Jorna JT (1985) Functions of the sliding pelvis in Xenopus laevis. Copeia

1985:251–254.

Vogt S, De Villiers FA, Ihlow F, et al (2017) Competition and feeding ecology in two sympatric Xenopus species (Anura: Pipidae). PeerJ 5:e3130. doi: 10.7717/peerj.3130 von Bertalanffy L (1938) A quantitative theory of organic growth (inquiries on grawth laws.

II). Hum Biol 10:181–213.

Weldon C, De Villiers A, Du Preez LH (2007) Quantification of the trade in Xenopus laevis from South Africa, with implications for biodiversity conservation. African J Herpetol 56:77–

83. References

White G, Burnham K (1999) Program MARK: survival estimation from populations of marked animals. Bird Study 46:120–139.

Whittingham MJ, Stephens PA, Bradbury RB, Freckleton RP (2006) Why do we still use stepwise modelling in ecology and behaviour? J Anim Ecol 75:1182–1189. doi:

10.1111/j.1365-2656.2006.01141.x

Wickham H (2009) ggplot2: elegant graphics for data analysis. Springer Science & Business

Media.

Wiens JJ, Graham CH (2005) Niche conservatism: Integrating evolution, ecology, and conservation biology. Annu Rev Ecol Evol Syst 36:519–539. doi:

10.1146/annurev.ecolsys.36.102803.095431

Wilcove DS, Rothstein D, Dubow J, et al (1998) Quantifying threats to imperiled species in the United States. Bioscience 48:607–615. doi: 10.2307/1313420

Wilson JRU, Dormontt EE, Prentis PJ, et al (2009) Something in the way you move: dispersal pathways affect invasion success. Trends Ecol Evol 24:136–144. doi:

10.1016/j.tree.2008.10.007

Wu Z-J, Li Y-M, Wang Y-P (2007) A comparison of stomach flush and stomach dissection in diet analysis of four frog species. Acta Zool Sin 53:364–372.

WWF (2014) Living Planet Report 2014: species and spaces, people and places. Gland,

Switzerland

207

Zug GR (1978) Anuran locomotion – structure and function, 2: jumping performance of semiaquatic, terrestrial, and arboreal frogs. Smithson Contrib to Zool 276:1–31. doi:

10.5479/si.00810282.276

References

209

APPENDICES

211

APPENDIX A

Courant J, Thirion J, Guillon M, et al (2014) Le régime

alimentaire de Xenopus laevis (Daudin, 1802) (Anura:

Pipidae) introduit en France. Bull la Société

Herpétologique Fr 150:1–7. Copy proofs: 15-05-2014

Bull. Soc. Herp. Fr. (2014) 150 : 1-7

Le régime alimentaire de Xenopus laevis (Daudin, 1802) (Anura : Pipidae) introduit en France par Julien Courant (1), Jean-Marc thirion (2), Michael Guillon (3), Pierre Grillet (4) & olivier Grosselet (5)

(1) 17, rue Marot 16250 Blanzac-Porcheresse [email protected] (2) Association Objectifs Biodiversités, 22 rue du Docteur Gilbert 17250 Pont l’Abbé d’Arnoult [email protected] (3) Centre d’Études Biologiques de Chizé 79360 Villiers-en-Bois, [email protected] (4) Nature Environnement Conseil 28, Place du 25 Août, 79340 Vasles [email protected] (5) Association Philofauna Ancienne École, 65120 Chèze [email protected]

Résumé – les espèces exotiques envahissantes, comme le Xénope lisse, Xenopus laevis (Daudin, 1802), présentent généralement un régime alimentaire susceptible d’impacter les communautés aquatiques qu’elles côtoient. la composition du régime alimentaire du Xénope lisse a déjà été étudiée sur son aire de répartition d’origine et dans plusieurs zones envahies par l’espèce. en France, cet aspect de l’écologie de Xenopus laevis n’a pour le moment jamais fait l’objet d’une étude. le présent article rap- porte les résultats obtenus à partir de dissections réalisées sur des individus provenant de mares des Deux-sèvres, à quelques kilomètres du point d’introduction de l’espèce sur le territoire français. le régime alimentaire comporte une importante variété de proies, avec une préférence marquée pour les proies d’origine aquatique. une analyse par calcul de fréquences d’occurrence a permis de mettre en évidence d’autres groupes, comme les adultes de lépidoptères, qui ont représenté une part importante des proies identiiées. Mots-clés : Xenopus laevis, régime alimentaire, espèce envahissante.

Summary – The diet of Xenopus laevis (Daudin, 1802) (Anura: Pipidae) introduced in France. invasive species like the african clawed frog, Xenopus laevis (Daudin, 1802) usually show a diet able to impact the local aquatic communities. the diet of Xenopus laevis has already been studied both in its original distribution area and in some invaded areas. in France, this side of the species ecology has not been studied yet. this article contains the results obtained from dissections realized on individuals from ponds of Deux-sèvres, at a few kilometres from the introduction point of the species. the diet shows an important variety of preys, with an unmistakable predilection for aquatic preys. an analysis by occurrence frequency has highlighted other groups, like Moths, which represented an important proportion of identiied preys. Key-words: Xenopus laevis, diet, invasive species.

- 1 - INTRODUCTION l’augmentation du nombre et l’expansion des espèces exotiques envahissantes repré- sentent une des causes du déclin des amphibiens dans le monde (alford & richards 1999). les mécanismes et l’intensité de ces phénomènes d’invasion varient selon les espèces et les écosystèmes dans lesquels elles sont introduites. Comprendre ces mécanismes représente un enjeu majeur pour les responsables des territoires et des espaces naturels sensibles ain de lut- ter eficacement contre l’expansion des espèces exotiques envahissantes. Les milieux aquatiques français souffrent de nombreuses introductions, d’espèces animales comme végétales, qui ont déjà causé d’importants dégâts dans les populations et les communautés d’espèces de ces milieux sensibles. Parmi les vertébrés exotiques envahissants, 21 espèces de Poissons, six espèces d’amphibiens, deux espèces de reptiles, trois espèces d’oiseaux et cinq espèces de Mammifères ont été dénombrés dans les zones humides de France (Pascal et al. 2006). en France, il existe plusieurs espèces exotiques qui sont naturalisées : la Grenouille verte des Balkans Pelophylax kurtmulleri (Gayda, 1940) dans le sud-est, le triton crêté ita- lien Triturus carnifex (laurenti, 1768) autour du lac léman, le sonneur à ventre de feu Bom- bina bombina (linnaeus, 1761) en lorraine, le Discoglosse peint Discoglossus pictus otth, 1837 dans le languedoc-roussillon, la Grenouille verte de Bedriaga Pelophylax bedriagae (Camerano, 1882) dans la vallée du rhône, la Grenouille verte de Berger Pelophylax bergeri (Günther in engelmann, Fritzsche, Günther & obst, 1986) en Corse et deux espèces qui pré- sentent un caractère envahissant : la Grenouille taureau, Lithobates (Aquarana) catesbeianus Dubois, 2006 dans trois régions et le Xénope lisse, Xenopus laevis dans le Centre-ouest (Pascal et al. 2006, Vacher 2008, thirion & evrard 2012). Ce dernier, originaire d’afrique subsaharienne, est présent naturellement du Cameroun jusqu’en afrique du sud. Cette espèce comportait six sous-espèces (Kobel et al. 1996), dont quatre ont depuis été élevées au statut d’espèce (Frost 2011). selon evans et al. (2011), une profonde révision des statuts des différentes lignées africaines de Xenopus laevis est encore nécessaire. le Xénope colonise plusieurs régions du monde, comme la Californie (McCoid & Fritts 1980), le Chili (lobos & Measey 2002), le royaume-uni (tinsley & McCoid 1996, Measey 1998), la sicile (lillo et al. 2005, Faraone et al. 2008) et le Portugal (rebelo et al. 2010). les populations établies de par le monde seraient originaires de la même région d’afrique du sud (Measey et al. 2012). l’espèce a longtemps été utilisée pour les tests de grossesse, puis comme modèle d’étude en laboratoire de recherche (tinsley et al. 1996), ce qui explique les importantes importations en provenance d’afrique du sud et des implantations locales suite à des lâchés volontaires ou accidentels. Pour se propager, l’espèce utilise principalement le réseau hydro- graphique ; elle se déplace dans une moindre mesure par des moyens terrestres (Fouquet & Measey 2006), notamment après les périodes de sécheresse pour migrer vers de nouveaux points d’eau (tinsley et al. 1996). a l’instar d’autres pays colonisés, la raison de son introduction en France est liée aux activités d’un laboratoire d’élevage de Xenopus laevis dans le nord des Deux-sèvres, qui fournissait des spécimens aux instituts de recherche. introduit en France dans un point d’eau unique du nord des Deux-sèvres dans les années 80 (thirion et al. 2009), le Xénope lisse serait actuellement répandue sur 207 km² (Measey et al. 2012). il a colonisé une partie du

- 2 - bassin versant de la loire (Fouquet & Measey 2006, Grosselet et al. 2006, Measey et al. 2012). une étude réalisée sur les populations françaises entre 2003 et 2005 a permis de faire un premier bilan concernant l’ampleur et le processus de la colonisation (Fouquet & Measey 2006, Grosselet et al. 2006) dans le but d’informer les responsables locaux et de déinir les méthodes de lutte possible contre Xenopus laevis (thirion et al. 2009). le régime alimentaire du Xénope lisse a fait l’objet de plusieurs études dans le monde, y compris dans la zone de répartition d’origine de l’espèce (schoonbee et al. 1992), mais aucune ne concerne la population française. Deux études ont été menées sur des populations introduites en Californie (McCoid & Fritts 1980) et au Pays de Galles (Measey 1998). elles ont notamment permis de mettre en évidence la diversité des proies capturées par Xenopus laevis, une préférence pour les taxons benthiques et le zooplancton et une stabilité du comportement alimentaire au cours des saisons (Ibid.). l’objectif de ce travail est de présenter le régime alimentaire du Xénope lisse dans une population introduite du nord des Deux-sèvres, à partir de l’analyse de contenus stomacaux d’individus capturés en 2005. Ces résultats pourront contribuer à l’évaluation de l’impact potentiel de cette espèce exotique envahissante sur les écosystèmes colonisés en France.

MATÉRIELS ET MÉTHODES Site d’étude

le groupe de mares (nmares = 4) qui ont été échantillonnées se situe autour de la com- mune de Bouillé-saint-Paul (79), non loin du point d’introduction de l’espèce. Ceci laisse supposer que le Xénope a colonisé ces mares dans les premières années de son invasion, il y a plus de vingt ans. Ces mares, spéciiques des milieux bocagers, présentent des caractéristi- ques environnementales biotiques et abiotiques analogues (thirion et al. 2009).

Animaux échantillonnés les individus ont été capturés au mois d’août 2005 dans des nasses plongées dans les quatre mares. Ces captures ont été réalisées durant deux journées consécutives (le 19 et 20 août). tous les individus capturés ont été sexés, mesurés et pesés et un échantillon choisi aléatoirement parmi ces individus (nalim = 23) a été sélectionné pour une analyse du contenu stomacal. les vingt-trois animaux capturés ont été euthanasiés (immersion dans un bain de chlo- rhydrate de benzocaïne dilué) et ont permis l’analyse des contenus stomacaux. la dissection et les analyses de contenus stomacaux ont eu lieu immédiatement après la mort des spéci- mens, pour éviter que la poursuite de la digestion n’altère l’identiication des proies.

Analyse des proies Chaque proie a été identiiée au niveau taxonomique de l’ordre sous une loupe bino- culaire (tab. i). une distinction entre les stades adultes et larvaires est réalisée lorsque les modes de vie de ces deux stades sont différents. le régime alimentaire a été décrit par la fréquence d’occurrence, l’occurrence relative et l’abondance relative des proies.

- 3 - • La fréquence d’occurrence, ou degré de présence (Lescure 1971) d’un type de proie correspond au rapport, exprimé en pourcentage, du nombre d’estomacs dans lesquels ce type de proie est présent par rapport à l’ensemble des estomacs analysés. • L’occurrence relative correspond au nombre d’individus d’une proie donnée par rapport à l’ensemble des types de proies trouvés au cours de l’analyse. • L’abondance relative traduit le pourcentage d’individus d’un type de proie par rapport au total des individus trouvés dans l’ensemble des estomacs analysés. le pourcentage de proies aquatiques a été comparé à celui des proies terrestres. Parmi les proies aquatiques, le pourcentage des espèces benthiques a été comparé à celui des espè- ces pélagiques. l’ensemble des calculs de comparaison ont été effectués avec le logiciel “r” ver. 2.12.0 (r Core team 2012) en utilisant des tests de Khi carré de conformité avec une correction de continuité de Yates.

RÉSULTATS sur les 23 Xénopes disséqués, cinq présentaient un estomac vide. l’analyse des résul- tats portera sur les contenus stomacaux des 18 individus restants (tab. i). les siluriformes détectés appartiennent aux stades juvéniles d’Ameuirus nebulosus (lesueur, 1819), le poisson-chat. une partie des larves de Diptères aquatiques ont été trouvées dans l’estomac du même individu, ce qui explique sa faible fréquence d’occurrence malgré un nombre important de spécimens. Aucun fragment de larve d’amphibien n’a été identiié dans l’échantillon

Tableau i : indices d’occurrence et d’abondance des différentes proies de Xenopus laevis. table i: occurrence and abundance indexes of the different Xenopus laevis preys.

Taxons et stades Nombre Fréquence Occurrence relative Abondance (N) d’occurrence (% sur 12 types de relative (% sur 18 estomacs) proies) (N / 38) x 100 Coléoptère aquatique 4 16,67 25,00 10,53 (adulte) Coléoptère aquatique 2 11,11 8,33 5,26 (larve) Diptère aquatique (larve) 9 16,67 8,33 23,68 Hétéroptère aquatique 4 16,67 25,00 10,53 (adulte) Coléoptère terrestre 2 11,11 16,67 5,26 (adulte) Diptère (adulte) 2 5,56 8,33 5,26 Orthoptère 1 5,56 8,33 2,63 Hyménoptère 1 5,56 8,33 2,63 Odonate (larve) 3 16,67 25,00 7,89 Lépidoptère (adulte) 6 22,22 33,33 15,79 Lépidoptère (larve) 1 5,56 16,67 2,63 Siluriforme 3 16,67 25,00 7,89 Total 38 ≈ 100,00

- 4 - de contenus stomacaux. seuls des adultes de Pelophylax sp. ont été observés dans les mares échantillonnées. les proies terrestres (Coléoptères, lépidoptères, orthoptères, hyménoptères, Diptères adultes) et les proies aquatiques (Coléoptères aquatiques, larves de Diptères, larves d’odonates, Hétéroptères et Siluriformes) ont été identiiées dans des quantités non statistiquement signiicativement différentes au seuil 5 % (χ² = 3,80, ddl = 1, p = 0,052). Au sein des proies aquatiques, les proies benthiques (larves de Diptères et d’odonates) et les proies pélagiques (Coléoptères aquatiques, hétéroptères, siluriformes) ne sont pas présentes en quantités statistiquement signiicativement différentes (χ² = 0,04, ddl = 1, p = 0,835) dans l’échantillon.

DISCUSSION lors de précédentes études sur le comportement alimentaire de Xenopus laevis en Cali- fornie (McCoid & Fritts 1980) et au Pays de Galles (Measey 1998), il a été révélé que celui- ci se nourrit potentiellement d’une grande variété de proies. Ces études ont pris en compte la quantité de proies, mais aussi la masse que celles-ci représentaient. Peu de travaux ont permis d’identiier des cas de prédations signiicatives de poissons. D’après lafferty et Page (1997), Xenopus laevis consomme de manière intensive Eucyclo- gobius newberryi (Girard, 1856), une espèce menacée aux États-unis. il s’agit à l’heure actuelle du seul cas avéré de menace sur une espèce indigène de poisson. Dans l’échantillon analysé sur la population française, seuls quelques poissons chat de petite taille, Ameuirus nebulosus ont été retrouvés. notre étude révèle que les proies dont le milieu de vie est aquatique sont majoritaires de manière non signiicative. Ce résultat peut s’expliquer par la faible taille de l’échantillon, entraînant un test statistique de puissance limitée. la prépondérance des proies aquatiques avait unanimement été constatée dans trois précédentes études menées au Pays de Galles (Measey 1998), en Californie (McCoid & Fritts 1980) et dans l’aire de répartition d’origine de l’espèce (schoonbee et al. 1992). D’après Measey (1998), les invertébrés terrestres sont minoritaires en nombre toute l’année, mais représentent une large proportion en termes de poids durant le printemps et l’été. il a été suggéré que Xenopus laevis emploie une méthode de chasse à l’affût qui le rend inadapté à la capture de proies nectoniques et à forte mobilité (avila & Frye 1978, McCoid & Fritts 1980). or, dans l’échantillon analysé, la plupart des proies recensées ont un caractère mobile et montrent une égale proportion de benthiques et de pélagiques. Ces résultats concor- dent avec ceux de Measey (1998) et suggèrent que le Xénope lisse possède les capacités pour la capture de proies situées dans le fond comme à la surface de l’eau. une étude approfondie en laboratoire permettrait de déinir précisément ces capacités ainsi que les préférences de capture, en agissant sur la biodisponibilité de chacun des types de proies. Les proies terrestres, qui représentent 34 % des proies identiiées, sont composées à presque 50 % de Lépidoptères adultes. Ces derniers présentent la plus forte fréquence d’occurrence de tout l’échantillon et la seconde abondance relative. tous les lépidoptères adultes, ainsi que la plupart des proies terrestres, sont issus de la même mare. il est probable que ce site était moins riche en invertébrés aquatiques que les autres, ce qui justiierait un tel résultat.

- 5 - les précédentes études menées sur le comportement alimentaire de Xenopus laevis n’ont pas révélé une telle représentation des lépidoptères parmi les proies terrestres. Ce résul- tat suggère que l’espèce possède un comportement alimentaire qui s’adapte aux ressources disponibles. les Chironomes, principaux représentants du groupe des Diptères dans l’échantillon, sont les proies retrouvées en plus grandes quantité parmi tous les groupes. Dans son analyse du régime alimentaire du Xénope lisse, Measey (1998) a démontré que l’espèce présente une préférence pour les larves et les pupes de Chironomes. Ces proies étant de faible masse, il est cependant dificile de connaître leur réelle contribution dans le régime alimentaire de Xeno- pus laevis. les échantillons ont été prélevés sur le terrain au cours du mois d’août. les points d’eau comportaient donc potentiellement des pontes ou des larves de Xenopus laevis. Cependant, aucun constat de cannibalisme n’a été répertorié dans aucune des quatre mares prospectées. D’après tinsley et al. (1996) et Measey (1998), les cas de cannibalisme sont surtout obser- vés durant la période d’accouplement. Ce comportement peut s’expliquer par l’augmentation des besoins en ressources nutritives de l’espèce à cette période de fortes activités (Measey, 1998). Chez cette espèce, la reproduction est déclenchée suite à une période de sécheresse et ou une raréfaction des ressources alimentaires (lodé, 2011). il est cependant probable que le comportement de cannibalisme soit sous-estimé dans l’échantillon analysé, comme cela a été suggéré lors de précédentes études (Measey 1998). D’après Crayon (2005), le cannibalisme sur les larves apparaît lors des périodes où les proies habituelles se raréient, et permet alors aux individus de se maintenir dans des points d’eau dont les faibles ressources devraient les forcer à émigrer. les résultats de cette étude préliminaire apportent les premières données sur le régi- me alimentaire de Xenopus laevis en France. ils ne permettent pas d’évaluer un impact sur la biodiversité, mais le caractère opportuniste de cet amphibien pourrait inclure des proies comme les larves et adultes d’amphibiens dans le régime alimentaire de l’espèce, notamment au printemps. Cet impact pourra être estimé par la comparaison des peuplements dans des sites colonisés et non colonisés, et par l’étude du régime alimentaire tout au long de l’année avec l’estimation des biomasses de proies potentielles dans les sites étudiés.

Remerciements – nous aimerions remercier le Conseil Général des Deux sèvres et la Dreal Poitou- Charentes pour leur soutien inancier ainsi que les agriculteurs du bocage pour leur accueil et leur inté- rêt porté à cette étude.

RÉFÉRENCES BIBLIOGRAPHIQUES alford r.a. & richards s.J. 1999 – Global amphibian declines: a problem in applied ecology. Ann. Rev. Ecol. Syst, 30: 133-165. avila V.l. & Frye P.G. 1978 – Feeding behaviour of the african clawed frog (Xenopus laevis, Daudin) (amphibia, anura, Pipidae); effect of prey type. J. Herpetol., 12: 391-396. Crayon J.J. 2005 – species account: Xenopus laevis. In lannoo M.J. (éd.), amphibians declines: status and Conservation of u.s. amphibians, Vol. 2. university of California Press, Berkeley: 522-525. evans B.J., Greenbaum e., Kusamba C., Carter t.F., tobias M.l., Mendel s.a. & Kelly D.B. 2011 – Description of a new octoploid frog species (anura: Pipidae: Xenopus) from the Democratic republic of the Congo, with a discussion of the biogeography of african clawed frogs in the albertine rift. J. Zool., Lond., 283: 276-290.

- 6 - Faraone F.P., lillo F., Giacalone G. & lo Valvo M. 2008 – the large invasive population of Xenopus laevis in sicily, italy. Amphibia-Reptilia, 29: 405-412. Ficetola G. F., thuiller W. & Miaud C. 2007 – Prediction and validation of the potential global distribution of a problematic alien invasive species – the american Bullfrog. Div. Distrib., 13: 476-485. Fouquet a. 2001 – Des clandestins aquatiques. Zamenis, 6: 10-11. Fouquet a. & Measey G.J. 2006 – Plotting the course of an african clawed frog invasion in Western France. Anim. Biol., 56: 95-102. Frost D.r. 2013 – amphibian species of the World: an online reference. Version 5.6. american Muse- um of natural history, new York, usa. Grosselet o., thirion J.-M., Grillet P. & Fouquet a. 2006 – le Xénope lisse, une nouvelle espèce invasive en France. Courr. Nat., 225: 22-27. Kobel h.r., loumont C. & tinsley r.C. 1996 – the extant species. In tinsley r.C. & Kobel h.r. (eds), the Biology of Xenopus. oxford university Press, oxford: 9-34. lafferty K. & Page C. 1997 – Predation on the endangered tidewater goby, Eucyclogobius newberyi, by the introduced african clawed frog, Xenopus laevis, with notes on the frog’s parasites. Copeia, 1997: 589-592. lescure J. 1971 – l’alimentation du crapaud Bufo regularis reuss et de la grenouille Discoglossus occipitalis (Günther) au sénégal. Bull. IFAN, 33a: 446-466. lillo F., Marrone F., sicilla a., Castelli G. & Zava B. 2005 – an invasive population of Xenopus laevis (Daudin, 1802) in italy. Herpetozoa, 18: 63-64. lobos G. & Measey G.J. 2002 – invasive populations of Xenopus laevis (Daudin) in Chile. Herpetol. J, 12: 163-168. lodé t. 2011 – la Guerre des sexes chez les animaux. odile Jacob, Paris. 331 p. McCoid M.J. & Fritts t.h. 1980 – notes on the diet of a feral population of Xenopus laevis (Pipidae) in California. Southwest. Nat, 25: 272-275. Measey G.J. 1998 – Diet of feral Xenopus laevis (Daudin) in south Wales, u.K. J. Zool., Lond., 246: 287-298. Measey G.J., rödder D., Green s.l., Kobayashi r., lillo F., lobos G., rebelo r. & thirion J.M. 2012 – ongoing invasions of the african clawed frog, Xenopus laevis: a global review. Biol. Invasions, 14: 2255-2270. Pascal M., lorvelec o. & Vigne J.-D. 2006 – invasions biologiques et extinctions. Édition Quae, Belin, Paris. 350 p. r Core team 2012 – r: a language and environment for statistical computing. r Foundation for statistical Computing, Vienna, austria. isBn 3-900051-07-0, url http://www.r-project.org/. rebelo r., amaral P., Bernardes M., oliveira J., Pinheiro P. & leitão D. 2010 – Xenopus laevis (Dau- din, 1802), a new exotic amphibian in Portugal. Biol. Invasions, 12: 3383-3387. schoonbee h. J., Prinsloo J. F. & nxiweni J. G. 1992 – observations on the feeding habits of larvae, juvenile and adult stages of african clawed frog, Xenopus laevis, in impoundments in transkei. Wat. S. Afr, 18: 227-236. tinsley r.C. & Mccoid M.J. 1996 – Feral populations of Xenopus outside africa. In tinsley r.C. & Kobel h.r. (eds), the Biology of Xenopus. oxford university Press, oxford: 81-94. tinsley r.C., loumont C. & Kobel h.r. 1996 – Geographical distribution and ecology. In tinsley r.C. & Kobel h.r. (eds), the Biology of Xenopus. oxford university Press, oxford: 35-59. thirion J.-M. & evrard P. 2012 – Guide des reptiles et amphibiens de France. Belin, Paris. 224 p. thirion J.-M., Doré F., Grillet P., Bitton G., Koch G. & Cotrel n. 2009 – Étude et propositions de modes de lutte contre le Xénope lisse, Xenopus laevis. Conseil général des Deux-sèvres. 59 p. Vacher J.-P. 2008 – Le sonneur à ventre jaune (Bombina variegata) Plan régional d’actions alsace 2012-2016. Dreal alsace, BuFo, strasbourg. 35 p. Manuscrit accepté le 27 janvier 2014

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APPENDIX B

Ihlow F, Courant J, Secondi J, et al (2016) Impacts of climate change on the global invasion potential of the African clawed frog Xenopus laevis. PLoS One 11:e0154869.

RESEARCH ARTICLE Impacts of Climate Change on the Global Invasion Potential of the African Clawed Frog Xenopus laevis

Flora Ihlow1*, Julien Courant2, Jean Secondi3,4, Anthony Herrel2, Rui Rebelo5,G. John Measey6, Francesco Lillo7, F. André De Villiers6, Solveig Vogt1, Charlotte De Busschere8, Thierry Backeljau8, Dennis Rödder1

1 Herpetological Section, Zoologisches Forschungsmuseum Alexander Koenig, Bonn, Germany, 2 UMR7179 CNRS/MNHN, Paris, France, 3 UMR CNRS 5023 LEHNA, University Lyon 1, Lyon, France, a11111 4 UMR CNRS 6554 LETG-LEESA, University of Angers, Angers, France, 5 Centre for Ecology, Evolution and Environmental Changes, Faculdade de Ciências da Universidade de Lisboa, Lisboa, Portugal, 6 Centre for Invasion Biology, Department of Botany & Zoology, Stellenbosch University, Private Bag X1, Matieland 7602, Stellenbosch, South Africa, 7 Dipartimento di Scienze e Tecnologie Biologiche Chimiche e Farmaceutiche, Università di Palermo, Palermo, Italy, 8 Royal Belgian Institute of Natural Sciences, Brussels, Belgium and University of Antwerp, Antwerp, Belgium

* [email protected] OPEN ACCESS

Citation: Ihlow F, Courant J, Secondi J, Herrel A, Rebelo R, Measey GJ, et al. (2016) Impacts of Abstract Climate Change on the Global Invasion Potential of the African Clawed Frog Xenopus laevis. PLoS ONE By altering or eliminating delicate ecological relationships, non-indigenous species are con- 11(6): e0154869. doi:10.1371/journal.pone.0154869 sidered a major threat to biodiversity, as well as a driver of environmental change. Global cli- Editor: Jake Kerby, University of South Dakota, mate change affects ecosystems and ecological communities, leading to changes in the UNITED STATES phenology, geographic ranges, or population abundance of several species. Thus, predicting Received: February 14, 2016 the impacts of global climate change on the current and future distribution of invasive species Accepted: April 20, 2016 is an important subject in macroecological studies. The African clawed frog (Xenopus laevis),

Published: June 1, 2016 native to South Africa, possesses a strong invasion potential and populations have become established in numerous countries across four continents. The global invasion potential of X. Copyright: © 2016 Ihlow et al. This is an open access article distributed under the terms of the laevis was assessed using correlative species distribution models (SDMs). SDMs were com- Creative Commons Attribution License, which permits puted based on a comprehensive set of occurrence records covering South Africa, North unrestricted use, distribution, and reproduction in any America, South America and Europe and a set of nine environmental predictors. Models were medium, provided the original author and source are built using both a maximum entropy model and an ensemble approach integrating eight algo- credited. rithms. The future occurrence probabilities for X. laevis were subsequently computed using Data Availability Statement: All relevant data are bioclimatic variables for 2070 following four different IPCC scenarios. Despite minor differ- within the paper and its Supporting Information files. ences between the statistical approaches, both SDMs predict the future potential distribution Funding: This research was funded by biodiversa of X. laevis, on a global scale, to decrease across all climate change scenarios. On a conti- (http://www.biodiversa.org/, project title: “Invasive biology of Xenopus laevis in Europe: ecology, impact nental scale, both SDMs predict decreasing potential distributions in the species’ native range and predictive models,” grant recipients: AH, RR, TB, in South Africa, as well as in the invaded areas in North and South America, and in Australia DR) and the Deutsche Forschungsgemeinschaft where the species has not been introduced. In contrast, both SDMs predict the potential (http://www.dfg.de/, DFG RO 4520/3-1, grant recipient: DR). The publication of this article was range size to expand in Europe. Our results suggest that all probability classes will be equally funded by the Open Access fund of the Leibniz affected by climate change. New regional conditions may promote new invasions or the Association. The funders had no role in study design, spread of established invasive populations, especially in France and Great Britain. data collection and analysis, decision to publish, or preparation of the manuscript.

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 1/19 Global Invasion Potential of Xenopus laevis

Competing Interests: The authors have declared Introduction that no competing interests exist. Plenty of evidence exists for impacts of climate change on ecosystems and ecological communities [1–8] Climate change modifies climatic factors such as ambient temperatures, precipitation, and the frequency of extreme weather events, which profoundly affects species’ geographical distributions [9]. Recent research reveals coherent patterns of ecological change across systems [10] concerning phenological changes, geographic range shifts and modifications in species abundance [11]. Rising ambient temperatures might promote range expansions beyond the northern range limits or favour altitudinal range shifts, while increasing temperatures enhance winter survival [9]. Non-indigenous species are expanding worldwide [12], and have been identified as a major driver of global biodiversity loss and environmental change [13–16]. Invasive species alter pro- ductivity, hydrology and nutrient cycles and thus, influence survival of native species and dis- rupt natural competition in host ecosystems [17]. Human-mediated transport for tourism or trade provides introduction pathways and therefore contributes to the rising number of introductions of alien species [18, 19]. Increases of biological invasions were found to coincide with the industrial revolution [18, 20, 21] and unprecedented acceleration of merchandise trade within the last 50 years led to progressive increases in the introduction of alien species [18]. Biological invasions are assumed to increase in the future in response to globalisation and climate change [18, 20–23]. Climate change is widely considered to exacerbate the impact of invasive species by making additional space suitable, enhancing survival and reproduction success, and by improving the competitive capacity of non-indigenous species [9, 24–26]. Depending on the physiological sensitivity to climatic conditions, the impact of climate change will vary among organisms (e.g. [27–29]) with poikilothermic taxa, such as amphibians, being particularly affected [11, 28, 30]. Due to behavioural traits, physiological processes and breeding phenology that closely depend on temperature and moisture, and a limited dispersal capacity, amphibians are especially sensitive and will be heavily affected by climate change [11, 31, 32]. However, climate change might also create opportunities for niche differentiation and evolution, by altering the composition of the resident biota, creating empty niches [33]. Fur- ther, impacts of climate change will likely vary geographically (e.g. [5, 11, 34, 35]). Although, considerable interest exists in predicting the spread and success of non-indigenous species, research linking climate change and biological invasions remain scarce [9, 36]. While the invasive potential of numerous species such as the invasive cane toad (Bufo marinus) in Australia [37, 38] was predicted to increase (e.g. [9, 39–41]) other studies suggest an opposite pattern e.g. for the American Bullfrog (Lithobates catesbeianus) in South America [42]. The African clawed frog, Xenopus laevis (Daudin, 1802), is one of the world's most widely distributed amphibians with populations originating from the Cape region in South Africa having become established on four continents (North America, South America, Asia and Europe) due to both accidental escape and voluntary release of laboratory animals [43–51]. While the establishment of introduced populations was most successful in areas with a Mediterranean climate, which closely resembles the environmental conditions of the Western Cape region, the persistence of several populations in cooler environments for decades suggests a capacity for long-term adaptation [52]. Recent research indicates that the global invasion potential of X. laevis has been severely underestimated with vast areas being potentially susceptible to invasion [50]. In addition, it has been claimed that climate change is likely to enhance this species’ invasion potential, favouring range expansion and population growth [53] particularly in Mediter- ranean climate regions.

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 2/19 Global Invasion Potential of Xenopus laevis

Macroecological approaches represent a popular analytical tool to assess and predict the impacts of climate change (e.g. [54–56]) as well as to assess spatial patterns of biological invasions in order to prioritize regions for the early detection of invasion outbreaks (e.g. [42, 57, 58]). The utility of Species Distribution Models (SDMs) to predict future spread of non-indigenous species has been demonstrated repeatedly (e.g. [57, 59, 60]). For general assumptions and methods see Elith and Lethwick [61]. Previous research targeting X. laevis showed large areas in Asia, southern Australia, south-western Europe, North and South America to be particularly vulnerable to colonization [52]. Owing to its ability to settle on various continents and its expected impact on local wetland communities [46, 49, 62–65], X. laevis is recognized as a major invasive amphibian species worldwide [50]. For such species it is crucial to develop accurate models of potential colonization and range shifts accounting for short-term climate changes. Such models may help to identify zones that could be colonized, refine risk assessments, and target prevention measures. Aiming to contribute to the future management of X. laevis, this study investigates the present and forecasts the future potential invasion range of the species on a global scale, based on an updated and extended occurrence data set and a similar set of bioclimatic variables as those used by Measey et al.[50]. We hypothesize that rising ambient temperatures associated with climate change will likely increase occurrence probabilities in currently cooler environments e.g. along the present day northern range limits of the species. Thus, climate change will promote the expansion of existing populations and increase the probability of additional invasions in the northern hemisphere, while occurrence probabilities for populations from the southernmost range limits are likely to decrease.

Materials and Methods In order to assess the present and future invasion potential of Xenopus laevis on a global scale georeferenced occurrence records, covering the species’ native distributional range in South Africa, as well as all known invasive populations in Europe, were obtained from recent literature [50] and supplemented by 286 new records collected during own field research (J.C., J.S., R.R., J.M., F.L., A.dV.). Until recently X. laevis was considered a species complex with a number of genetically distinct lineages [66]. Measey et al.[50] noted that all invasive populations were from the South African ‘Cape’ clade [67, 68] an area including the winter rainfall region and southern coast of South Africa. Since that time, De Busschere et al.[69] determined that the French invasion incorporated lineages from throughout the range of X. laevis. In addition, recent phylogeographic research supports the recognition of X. petersii, X. victorianus and X. poweri as separate species and therefore confines the native range of X. laevis sensu stricto to southern Africa (South Africa, Lesotho, Swaziland, Namibia, Botswana, Zimbabwe, Mozam- bique and Malawi) [70]. The occurrence data set for southern Africa used in this study was adjusted accordingly. It is important to note that this is a novel interpretation of the taxonomy and native range of X. laevis in comparison to that used by Measey et al.[50]. To prevent over‐fit and false inflation of model performance through spatially auto‐correlated species records [71–73], the comprehensive dataset of 1382 records was filtered, and clustered locality records were reduced to a single point within a specified Euclidian distance in environmental space using the spatially rarefy occurrence data tool for the ArcGIS SDM toolbox [74]. The final dataset used to build SDMs contained 826 records for South Africa, 37 for South America, 24 for North America, and 38 for Europe. SDMs built exclusively on occurrence data from the native range of invasive species tend to underestimate the potential invasive distribution, particularly when projecting onto climate change scenarios [75]. Hence, native and invasive occurrence records were pooled [50, 76] for the computation of SDMs. In this way a maximum amount of information on the species’ realized bioclimatic niche was integrated.

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As environmental predictors, 19 bioclimatic variables available through the WorldClim- database ([77] http://www.worldclim.org/bioclim) were used. They represent minima, maxima and average values of monthly, quarterly, and annual ambient temperature as well as precipitation recorded between 1950 and 2000. All predictors had a spatial resolution of 2.5 arc min (approx. ~5 km resolution at the equator). Out of the total set of variables a set of nine predictors with pairwise Spearman rank correlation coefficients R2 < 0.75 were selected to minimize predictor correlation. Subsequently, SDMs were computed using the machine learning algorithm Maxent version 3.3.3k [78, 79]. Maxent is supposed to exhibit a higher predictive performance than more con- ventional techniques [61] and has successfully been used to model the potential distribution of invasive species and to assess impacts of climate change [78]. Only linear, quadratic and prod- uct features were allowed in order to restrict model complexity, while extrapolation was not allowed to reduce uncertainties due to projections onto non-analogous climates [80, 81]. The Maxent model was trained by randomly splitting the species records into 70% used for model training and 30% for model testing applying a bootstrap approach. The Area Under the receiver operating characteristic Curve (AUC) [82] was used to evaluate the discrimination ability of the resulting SDM. Averages across 100 replicates were used for further processing. As the selection of an appropriate background is known to affect model performance [83], a circular buffer of 250 km around each locality record was selected as training area following Measey et al.[50]. Ensemble SDMs were computed using the biomod2 package version 3.2.2 [84] for Cran R [85] including the following eight algorithms: Generalized Linear Models (GLM), Generalized Boosting Models (GBM), Generalized Additive Models (GAM), Classification Tree Analysis (CTA), Artificial Neural Networks (ANN), Factorial Discriminant Analysis (FDA), Maxent, and Multivariate Adaptive Regression Splines (MARS) applying the same training background as for the Maxent analyses. We applied a bootstrapping approach with 100 replicates randomly subdividing the locality dataset in 70% for model training, whereas the remaining 30% were used for model evaluation using the AUC [82], True Skill Statistic (TSS) and Cohen’s Kappa [86]. The average projection across all replicates was used for further processing. The “Mini- mum training presence” (mtp) referring to the lowest generated probability estimate of the training data [87], was applied as presence-absence threshold. While threshold selection is a potential source for biases, the mtp has been shown to represent a confident method perform- ing well for presence only SDMs [88] particularly for modelling potential distributions of invasive species [88, 89]. Areas requiring extrapolation beyond the training range of the variables were quantified using a multivariate environmental similarity surface (MESS) analysis [90] for Maxent and conceptually equivalent clamping masks for biomod2. To predict the future potential distribution of X. laevis on a global scale, 11 general circulation models (GCMs: BCC-CSM1-1, CCSM4, GISS-E2-R, HadGEM2-AO, hadGEM2-ES, IPSL-CM5A-LR, MIROC-ESM-CHEM, MIROC-ESM, MIROC5, MRI-CGCM3 and Nor- ESM1-M) representing simulations for four representative concentration pathways (RCP2.6, RCP4.5, RCP6, RCP8.5) for 2070 were obtained from the fifth assessment of the Intergovern- mental Panel for Climate Change (IPCC AR5 WG1 2013; http://www.ipcc.ch,[91]). The selected RCPs represent four possible greenhouse gas emission trajectories ranging from low (RCP2.6) to high (RCP8.5) corresponding to increases in global radiative forcing, from pre- industrial times to 2100. These climate projections were statistically downscaled to match the bioclim variables using the delta method [77], (http://www.worldclim.org/downscaling) for details also see [92,93]. As differences between the selected GCMs might cause uncertainty in

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SDM projections [59], average values across all GCMs were calculated for each RCP respectively. Finally, SDMs were projected onto the derived future climate change scenarios. In order to quantify impacts of different RCP scenarios onto the global invasion potential of X. laevis, the following predicted areas were determined: a) the entire SDM area using the ‘minimum training presence’ as threshold, b) the respective MESS area, and c) the SDM area– MESS area. A probability cut-off of 25%, 50%, and 75% onto the ‘SDM area–MESS area’ was applied to assess impacts for different probability classes. Further, the full model was parti- tioned into estimates for each continent. Subsequently, the invasion potential for X. laevis was determined on a continental scale as described above. Shift maps were generated to illustrate predicted gains, losses and stability of environmentally suitable space for all climate change scenarios following Bertelsmeier et al.[94]. Comparisons between the results obtained by our Maxent analyses, the results presented by Measey et al.[50] and the results obtained via bio- mod2 were performed by rescaling the probability output and subtracting the potential distribution grids from each other. Rescaling involved subtracting the minimum training presence threshold from each model and computation of percentages per grid cell relative to the maximum probability. The resulting maps quantitatively indicate for each area which SDM shows a higher probability.

Results

Model performance was 0.841 (AUCtest) and 0.846 (AUCtraining) for the maximum entropy model while weighted means of 0.631 for Cohen’s Kappa, 0.892 for AUC, and 0.685 for TSS were obtained for the ensemble model, demonstrating that both SDMs discriminate moder- ately well between suitable versus unsuitable space [79]. For the maximum entropy model the contribution of eight predictors exceeded 5%, while for the ensemble approach seven predictors had a contribution exceeding 5% (Table 1, S1 Table). Predictor contribution for the maximum entropy model was particularly high for ‘precipitation of the driest quarter’ (27.65%), ‘mean temperature of the wettest quarter’ (16.82%), ‘mean temperature of the coldest quarter’ (14.52%), and ‘precipitation of the warmest quarter’ (11.38%) (Table 1). Variables with high contribution in the ensemble model were ‘mean temperature of the coldest quarter’ (19.05%), ‘precipitation of the warmest quarter’ (16.57%), ‘mean temperature of the warmest quarter’ (13.92%), ‘precipitation of the driest quarter’ (12.56%) (Table 1). For the respective response curves see S1 Fig & S2 Fig. Both modelling approaches yielded similar global patterns for the present potential distribution of X. laevis. However, the more complex ensemble SDM predicted larger areas with

Table 1. Variable contribution for the maximum entropy and the ensemble SDM.

Variable Contribution (%)

ID Bioclimatic Variable Maxent SDM Ensemble Bio 17 precipitation of driest quarter 27.65 12.56 Bio 8 mean temperature of the wettest quarter 16.82 7.61 Bio 11 mean temperature of coldest quarter 14.52 19.05 Bio 18 precipitation of warmest quarter 11.38 16.57 Bio 19 precipitation of coldest quarter 8.25 8.25 Bio 7 temperature annual range 6.99 4.96 Bio 9 mean temperature of driest quarter 6.24 8.33 Bio 16 precipitation of wettest quarter 6.21 7.95 Bio 10 mean temperature of warmest quarter 1.93 13.92 doi:10.1371/journal.pone.0154869.t001

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Fig 1. Global projection of the potential distribution of X. laevis A) derived from the maximum entropy SDM; B) derived from the ensemble SDM. Probability ranging from moderate (dark blue) to highly suitable (yellow). doi:10.1371/journal.pone.0154869.g001

slightly higher probabilities than the maximum entropy SDM (Fig 1 & S3 Fig). While the ensemble SDM predicted 28% of the world’s surface to be presently environmentally suitable for X. laevis, the maximum entropy SDM predicted only 12% (Table 2). As expected, high occurrence probability was predominantly predicted for areas resembling climatic characteristics of the South African Cape region with high annual variation in ambient temperatures, comparatively warm and dry summers, and mild and wet winters (S1 Fig & S2 Fig). More precisely, probability was correlated with mild winter temperatures (10–20°C), low precipitation during the wettest (<350 mm), and the coldest quarter (<500 mm) and high precipitation during the warmest quarter (400–600 mm) (S1 Fig & S2 Fig).

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Table 2. Environmentally suitable space given as percent of the world’s surface area for current climatic conditions and projections onto climate change scenarios. Percentages refer to the SDM-MESS area; values increasing with climate change scenarios are displayed in bold.

Maximum entropy SDM Continent % current % RCP 2.5 % RCP 4.5 % RCP 6 % RCP 8.5 Africa 9 6 5 5 3 Europe 4 8 9 10 11 North America 10 8 7 7 7 South America 26 23 21 21 20 Australia 40 33 31 30 28 Asia 8 9 10 10 11 global 12 12 11 11 11 Ensemble SDM % current % RCP 2.5 % RCP 4.5 % RCP 6 % RCP 8.5 Africa 24 16 12 12 9 Europe 21 30 33 35 38 North America 24 29 31 30 30 South America 30 31 28 28 26 Australia 74 72 66 66 56 Asia 24 20 19 19 18 global 28 27 26 26 25 doi:10.1371/journal.pone.0154869.t002

According to both models, there are regions exhibiting high occurrence probabilities under current climatic conditions located on all continents (Fig 1), but percentages of environmentally suitable space varied greatly (Table 2). Both SDMs predict only moderate coverage with environmentally suitable space for X. laevis to presently exist in the northern hemisphere, while high coverage was predicted for Australia and South America (Table 2). Potentially highly suitable areas cover the south central United States (Texas, Kansas), western (stretching from Peru through Colombia and into Venezuela), eastern (eastern Brazil, stretching along the Paraná River), and southern South America (north-eastern Argentina). In Europe, particularly high probabilities are predicted in Portugal, eastern Spain, southern France, and Italy. In accordance with Measey et al.[50], both SDM approaches predict only moderate occurrence probability in Great Britain. In addition to the native distribution of X. laevis covering vast areas in southern Africa, there is a high occurrence probability in Morocco and the eastern Afromon- tane region (where no invasions by X. laevis have been reported so far). Moreover, both SDMs predict high suitability in China (Nanzhao plateau, Sichuan, Chinese plain), Japan, and southern Australia. The maximum entropy SDM highlights additional regions with high probabilities in Florida and south-eastern China (Zhejiang Province) (Fig 1, S3 Fig). On a global scale, both SDM approaches predict suitable range sizes for X. laevis to decrease across all four RCP scenarios (Table 2, Fig 2 & Fig 3). However, the magnitude of decrease var- ies between RCP scenarios and between SDM approaches. For the maximum entropy SDM, the potentially suitable range size is predicted to decrease by 7 to 13% from RCP 2.5 to RCP 8.5, respectively. This corresponds to a maximum decrease of 1% in relation to the world’s surface area (Table 2). The ensemble SDM predicts a rather moderate decrease of 1 to 10%. Since the areas predicted by the ensemble SDM are generally larger, the portion of the world’s surface area predicted to be suitable shrinks by 3% (Table 2). A comparison between the full model and the respective probability cut-offs does not reveal significant differences between RCP scenarios (Fig 4), suggesting that probability classes will be equally affected by climate change.

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Fig 2. Global shift maps derived from Maxent illustrating predicted gains (dark violet) and losses (dark blue) of environmentally suitable space for different climate change scenarios; A) IPCC RCP2.6, B) IPCC RCP4.5, C) IPCC RCP6, D) IPCC RCP8.5. doi:10.1371/journal.pone.0154869.g002

On a continental scale, both SDM approaches suggest range sizes to decrease for the species’ native range in South Africa, as well as for the invaded areas in South America and Australia, where the species has not been introduced (Fig 5, Table 2, S2 Table & S3 Table). In contrast, the potential range in Europe will expand in response to climate change (Table 2). Concor- dantly, shift maps highlight different magnitudes of expected gains in Europe (Fig 2 & Fig 3). However, there are minor differences between the results of both approaches: while the maximum entropy SDM predicts range sizes to decrease for all probability classes, the ensemble SDM suggests range sizes to increase for North America when applying the mtp threshold (Fig 5). Furthermore, the ensemble SDM predicts increasing range sizes for the >25% probability class in Asia (Fig 5), while the maximum entropy model predicts increases for the mtp threshold, as well as for the probability classes >25% and >50% (Fig 5). Finally, the ensemble SDM predicts increasing range sizes only for the mtp threshold, while the maximum entropy SDM suggests an area expansion across all probability classes in Europe (Fig 5, S2 Table & S3 Table).

Discussion Our predictions of the present potential distribution of Xenopus laevis generally agree with the findings presented by Measey et al.[50] highlighting analogous spatial extents and geographic locations. Both SDMs reveal large areas that can potentially be colonized on several continents. However, the overlap comparison of our SDMs with the prediction by Measey et al.[50] yields higher probability values on all continents for both our ensemble and our maximum entropy SDM (S3 Fig). These differences might be attributed to the different occurrence data sets used. While Measey et al.[50] applied a restricted definition within X. laevis using only records from a single clade occurring in the winter rainfall zone and southern coast of South Africa [66], the data used in this study was adjusted according to the findings of Furman et al.[70] resulting in a larger coverage for the species native range. Thus, the environmental space covered by the occurrence dataset used was larger than in Measey et al.[50]. Under current climatic conditions, both the Maxent and the ensemble model identify regions with suitable climatic conditions favouring invasion in Portugal, France, Sicily, Califor- nia, Chile, and Japan, where invasive populations already exist. Both SDMs predict only moderate probability for Great Britain, where populations from Wales and Lincolnshire have recently been extirpated [53]. Furthermore, our results highlight areas in Spain (including the Balearic Islands), mainland Italy (including Sardinia), and southern France (including Corsica) to be highly vulnerable to potential invasions, as these regions exhibit suitable climatic conditions for X. laevis and are adjacent to established invasive populations. Globally, the same applies to Baja California and central Mexico. Future projections of both SDM approaches identify regions that will likely become vulnerable to colonization in response to climate change. With an expected decrease of 1–3% (given as percentage of the worlds’ surface area) the overall magnitude of expected changes appears to be moderate, while the predicted global area suitable for X. laevis remains stable or slightly decreases with increasing RCP scenarios. However, predictions for Europe are the major excep- tion to this general trend with particularly good prospects for the invasive populations in north-western Europe (Figs 2 & 3). Xenopus laevis is capable of enduring extreme conditions ([50] and references therein). However, reproduction seems to be triggered by rainfall and

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Fig 3. Global shift maps derived from the ensemble SDM illustrating predicted gains (dark violet) and losses (dark blue) of environmentally suitable space for different climate change scenarios; A) IPCC RCP2.6, B) IPCC RCP4.5, C) IPCC RCP6, D) IPCC RCP8.5. doi:10.1371/journal.pone.0154869.g003

increasing temperatures [95, 96]. These physiological restrictions are well reflected in the variable contributions of the models. These were highest in the precipitation of driest and warmest quarter and in the temperature of the wettest, coldest and warmest quarters affecting reproduction cycles. While reproduction occurs throughout the year in California [97] the lower lethal limit of temperature tolerance in embryos was reported to be 10°C [95]. Harsh conditions only

Fig 4. Predicted development of area sizes suitable for X. laevis on a global scale; a) for the Maximum entropy SDM and; b) for the ensemble SDM. Mtp = minimum training presence, all areas sizes refer to SDM area–MESS area. doi:10.1371/journal.pone.0154869.g004

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Fig 5. Predicted development of area sizes suitable for X. laevis on a continental scale; left) for the Maximum entropy SDM, right) for the ensemble SDM. Mtp = minimum training presence, all areas sizes refer to SDM area–MESS area. doi:10.1371/journal.pone.0154869.g005

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permitted infrequent reproduction in Great Britain [53] and rising temperatures will likely improve physiological performances, fecundity, breeding success, and increase the rate of larval development. Thus, regional patterns may facilitate new invasions or promote a spread of the established invasive populations, especially in France and Great Britain, where populations persisted for decades [52]. While invasive populations are already spreading in France, British populations are presently considered extirpated [53]. Due to an increased environmental suitability caused by climate change along the northernmost boundaries of the species’ range, chances of successful establishment in Great Britain in case of re-introductions will increase in the future. Climate change is widely considered to exacerbate the impact of invasive species [9, 25] enhancing the invasive potential of some species (e.g. [9, 39–41]), including the invasive cane toad (Rhinella marina) in Australia [37, 38]. However, some studies suggest an opposite pattern e.g. for the global invasion potential of an assemblage of ant species [98], or the American Bullfrog (Litho- bates catesbeianus) in South America [42]. For X. laevis, it has been claimed that climate change will likely favour range expansion and population growth [53] particularly in Mediterranean climate regions. While on a global scale our predictions reveal the species’ potential range to decrease in response to climate change, populations from the northern hemisphere are predicted to expand. As X. laevis is kept as a model organism in laboratories all across the world and is still traded intensively [51, 98] Measey et al.[51] emphasize the importance of biosecurity at breeding facilities to prevent further escape and voluntary release of frogs and tadpoles [99]. While by now scientists working with X. laevis are likely to be aware of the species’ invasion potential [100], this frog is also readily available via the pet trade [98]. Once introduced, the species rapidly disperses using irrigation canals, ponds, and rivers as migration corridors, but also per- forms terrestrial migrations [101] even without rainfall [102]. Estimated annual spread of feral populations varied between 1 km [101] in France and 5.4 km [62] in Chile. Recent research found the maximum overland dispersal in native populations to be 2.3 km (Euclidean distance) within 6 weeks [102]. Although invasive X. laevis have demonstrably negative impacts on resident amphibian and fish communities [49, 63, 64, 103], attempts to eliminate invasive populations are limited. Recent studies stressed the urgent need for rigorous and comprehensive invasive species risk assessments to contribute to the development of management strategies [104, 105]. Prevention is generally considered more effective and cheaper than control and eradication of established populations [8, 25]. SDMs represent a quick and cost efficient tool to evaluate the current and future invasion potential of non-indigenous species. In addition, SDMs facilitate the identification of areas with high susceptibility to invasion and help to prioritize management actions. According to our results preventive measures should predominantly focus on the species’ northern range limits, particularly north-western Europe to prevent further spread and establishment of new populations as well as a re-introduction in Great Britain. Even though considered particularly difficult in Mediterranean areas, eradication of invasive populations of X. laevis has been proposed [101] and an eradication program was established by the Portuguese Governmental Nature Conservation Institute in Oeiras, western Portugal in 2010 [50]. In addition, eradication was successfully executed in Scunthorpe, Humberside area, Great Britain [50]. According to our results eradication of established populations of X. laevis should focus on areas where populations are still small and scattered, but likely to expand in response to climate change e.g. Portugal and Sicily.

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Supporting Information S1 Fig. Maximum entropy SDM response curves for selected predictor variables displaying relationships between predictor variables and occurrence probability of Xenopus laevis. Model contribution was assessed by building the model using a single corresponding predictor variable. The logistic output (probability of presence) is displayed on the Y axis. Red curves refer to mean responses of 100 replicate Maxent runs while the mean +/- one standard deviation is displayed in blue. (PDF) S2 Fig. Response curves for ensemble SDM showing the relationships between environmental predictors and occurrence probability of Xenopus laevis. The logistic output (probability of occurrence) is displayed on the Y axis. (PDF) S3 Fig. A) Overlap analysis of the ensemble SDM and the maximum entropy SDM, with red highlighting areas where the ensemble SDM yields higher probabilities and blue depicting areas where the maximum entropy SDM yields higher probabilities; B) Overlap analysis of the ensemble SDM and the SDM by Measey et al. (2012), with red indicating higher probability of the ensemble SDM and blue showing higher values for the SDM by Measey et al. (2012); C) Overlap analysis of the maximum entropy SDM and the SDM by Measey et al. (2012), with red highlighting regions with higher probability of the ensemble SDM and blue showing higher probability values for the SDM by Measey et al. (2012). Colour saturation increases with deviation of the models. Areas where both SDMs yield similar probability values are displayed in white. (PDF) S1 Table. Variable contribution for single algorithms of the ensemble SDM. (XLSX) S2 Table. Predicted area sizes in mio. km² for the maximum entropy SDM; maximum values for each threshold (mtp, 25%, 50%, 75%) are displayed in bold. (XLSX) S3 Table. Predicted area sizes in mio. km² for the ensemble SDM; maximum values for each threshold (mtp, 25%, 50%, 75%) are displayed in bold. (XLSX)

Author Contributions Conceived and designed the experiments: FI DR. Performed the experiments: FI DR. Analyzed the data: FI DR. Contributed reagents/materials/analysis tools: JC JS RR GJM FL FADV. Wrote the paper: FI DR JC JS AH RR GJM FL FADV SV CDB TB.

References 1. Hoegh-Guldberg O. Climate change, coral bleaching and the future of the world’s coral reefs. Mar Freshw Res. 1999; 50: 839–866. 2. Smith RC, Ainley D, Baker K, Domack E, Emslie S, et al. Marine Ecosystem Sensitivity to Climate Change Historical observations and paleoecological records reveal ecological transitions in the Ant- arctic Peninsula region. BioScience. 1999; 49: 393–404. 3. Parmesan C, Root TL, Willig M. Impacts of extreme weather and climate on terrestrial biota. Bull Am Meteorol Soc. 2000; 81: 443–450. 4. Peñuelas J, Flella I. Responses to a warming world. Science. 2001; 294: 793–795. PMID: 11679652

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 14 / 19 Global Invasion Potential of Xenopus laevis

5. Pounds JA. Climate and amphibian declines. Nature. 2001; 410: 639–664. PMID: 11287931 6. Walther GR, Post E, Convey P, Menzel A, Parmesan C, Beebee TJC, Fromentin JM, Hoegh- Gulberg O, Bairlein F. Ecological responses to recent climate change. Nature. 2002; 416: 389–395. PMID: 11919621 7. Thomas CD, Cameron A, Green RE, Bakkenes M, Beaumont LJ, et al. Extinction risk from climate change. Nature. 2004; 427: 145–148. PMID: 14712274 8. Simberloff D, Martin J-L, Genovesi P, Maris V, Wardle DA, et al. Impacts of biological invasions— what’s what and the way forward. Trends Ecol Evol. 2013; 28: 58–66. doi: 10.1016/j.tree.2012.07.013 PMID: 22889499 9. Dukes JS, Mooney HA. Does global change increase the success of biological invaders?. Trends Ecol Evol. 1999; 14: 135–139. PMID: 10322518 10. Walther GR, Post E, Convey P, Menzel A, Parmesan C, Beebee TJC, Fromentin JM, Hoegh- Gulberg O, Bairlein F. Ecological responses to recent climate change. Nature. 2002; 416: 389–395. PMID: 11919621 11. Parmesan C, Yohe G. A globally coherent fingerprint of climate change impacts across natural systems. Nature. 2003; 421: 37–42. PMID: 12511946 12. Sala OE, Chapin FS, Armesto JJ, Berlow E, Bloomfield J, et al. Biodiversity scenarios for the year 2100. Science. 2000; 287: 1770–1774. PMID: 10710299 13. Mack RN, Simberloff D, Lonsdale WM, Evans H, Clout M, Bazzaz FA. Biotic invasions: causes, epidemiology, global consequences, and control. Ecol Appl. 2000; 10: 689–710. 14. Lodge DM. Responses of lake biodiversity to global changes. In: Future scenarios of global biodiversity. Chapin FS III, Sala OE, Huber-Sannwald E, editors. Springer: New York; 2001. pp. 277–312. 15. Clavero M, García-Berthou E. Homogenization dynamics and introduction routes of invasive freshwater fish in the Iberian Peninsula. Ecol Appl. 2006; 16: 2313–2324. PMID: 17205906 16. Bálint M, Domisch S, Engelhardt CHM, Haase P, Lehrian S, Sauer J, Theissinger K, Pauls SU, Nowak C. Cryptic biodiversity loss linked to global climate change. Nature Clim Chang. 2011; 1: 313– 318. 17. Cramer W, Bondeau A, Woodward FI, Prentice IC, Betts RA, Brovkin V, Cox PM, Fisher V, Foley JA, Friend AD. Global response of terrestrial ecosystem structure and function to CO2 and climate change: results from six dynamic global vegetation models. Glob Chang Biol. 2001; 7: 357–373. 18. Hulme PE. Trade, transport and trouble: managing invasive species pathways in an era of globalization. J Appl Ecol. 2009; 46: 10–18. 19. Butchart S, Walpole M, Collen B, van Strien A, Scharlemann JPW, et al. Global biodiversity: indicators of recent declines. Science. 2010; 328: 1164–1168. doi: 10.1126/science.1187512 PMID: 20430971 20. Perrings C, Dehnen-Schmutz K, Touza J, Williamson M. How to manage biological invasions under globalization. Trends Ecol Evol. 2005; 20: 212–215. PMID: 16701371 21. Meyerson LA, Mooney HA. Invasive alien species in an era of globalization. Front Ecol Environ. 2007; 5: 199–208. 22. Lockwood JL, Cassey P, Blackburn T. The role of propagule pressure in explaining species invasions. Trends Ecol Evol. 2005; 20: 223–228. PMID: 16701373 23. McNeely JA. As the world gets smaller, the chances of invasion grow. Euphytica. 2006; 148: 5–15. 24. Corn PS. Climate change and amphibians. Anim Biodiv Conserv. 2005; 28: 59–67. 25. Pyke CR, Thomas R, Porter RD, Hellmann JJ, Dukes JS, Lodge DM, Chavarria G. Current practices and future opportunities for policy on climate change and invasive species. Conserv Biol. 2008; 22: 585–592. doi: 10.1111/j.1523-1739.2008.00956.x PMID: 18577088 26. Walther GR, Roques A, Hulme PE, Sykes MT, Pyšek P, Kühn I, Zobel M, Bacher S, Botta-Dukát Z, Bugmann H. Alien species in a warmer world: risks and opportunities. Trends Ecol Evol. 2009; 24: 686–693. doi: 10.1016/j.tree.2009.06.008 PMID: 19712994 27. Deutsch CA, Tewksbury JJ, Huey RB, Sheldon KS, Ghalambor CK, Haak DC, Martin PR. Impacts of climate warming on terrestrial ectotherms across latitude. Proc Natl Acad Sci USA. 2008; 105: 6668– 6672. doi: 10.1073/pnas.0709472105 PMID: 18458348 28. Sinervo B, Mendez-De-La-Cruz F, Miles DB, Heulin B, Bastiaans E, et al. Erosion of lizard diversity by climate change and altered thermal niches. Science. 2010; 328: 894–899. doi: 10.1126/science. 1184695 PMID: 20466932 29. Rowland EL, Davison JE, Graumlich LJ. Approaches to evaluating climate change impacts on species: a guide to initiating the adaptation planning process. Environ. Manage. 2011; 47: 322–337. doi: 10.1007/s00267-010-9608-x PMID: 21259061

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 15 / 19 Global Invasion Potential of Xenopus laevis

30. Miller D, Summers J, Silber S. 2004. Environmental versus genetic sex determination: a possible factor in dinosaur extinction? Fertil Steril. 2004; 81: 954–964. PMID: 15066448 31. Blaustein AR, Belden LK, Olson DH, Green DM, Root TL, Kiesecker JM. Amphibian breeding and climate change. Conserv Biol. 2001; 15: 1804–1809. 32. Beebee TJ. Amphibian breeding and climate. Nature. 2009; 374: 219–220. 33. Ackerly DD. Community assembly, niche conservatism, and adaptive evolution in changing environments. Int J Plant Sci. 2003; 164: 165–184. 34. Beaumont LJ, Pitman A, Perkins S, Zimmermann NE, Yoccoz NG, Thuiller W. Impacts of climate change on the world’s most exceptional ecoregions. Proc Natl Acad Sci USA. 2011; 108: 2306–2311. doi: 10.1073/pnas.1007217108 PMID: 21262825 35. Pounds JA, Fogden, MP, Campbell JH. 1999. Biological response to climate change on a tropical mountain. Nature. 1999; 398: 611–615. 36. Stachowicz JJ, Terwin JR, Whitlatch RB, Osman RW. Linking climate change and biological invasions: ocean warming facilitates nonindigenous species invasions. Proc Natl Acad Sci USA. 2002; 99: 15497–15500. PMID: 12422019 37. Urban MC, Phillips BL, Skelly DK, Shine R. The cane toad's (Chaunus (Bufo) marinus) increasing ability to invade Australia is revealed by a dynamically updated range model. Proc R Soc B. 2007; 274: 1413–1419. PMID: 17389221 38. Sutherst RW, Floyd RB, Maywald GF. The potential geographical distribution of the cane toad, Bufo marinus L. in Australia. Conserv Biol. 1995; 9: 294–299. 39. Kriticos DJ, Sutherst RW, Brown JR, Adkins SW, Maywald GF. Climate change and the potential distribution of an invasive alien plant: Acacia nilotica ssp. indica in Australia. J Appl Ecol. 2003; 40: 111– 124. 40. Thuiller W, Richardson DM, Midgley GF. Will climate change promote alien plant invasions? In: Bio- logical invasions. Springer, Berlin Heidelberg; 2007. Pp. 197–211. 41. Vilà M, Corbin JD, Dukes JS, Pino J, Smith SD. Linking plant invasions to global environmental change. In: Terrestrial ecosystems in a changing world. Springer, Berlin Heidelberg; 2007. Pp. 93– 102. 42. Nori J, Urbina-Cardona JN, Loyola RD, Lescano JN, Leynaud GC. Climate change and American Bullfrog invasion: what could we expect in South America. PloS ONE 2011; 6: e25718. DOId: doi: 10. 1371/journal.pone.0025718 PMID: 21991339 43. McCoid MJ, Fritts TH. Female reproductive potential and winter growth of African Clawed frogs (Pipi- dae: Xenopus laevis) in California. California Fish and Game 1995; 81: 39–42. 44. Tinsley RC, McCoid MJ. Feral populations of Xenopus outside Africa. In: Tinsley RC, Kobel HR, editors. The Biology of Xenopus. Oxford University Press, Oxford; 1996. Pp. 81–94. 45. Fouquet A. 2001. Des clandestins aquatiques. Zamenis. 2001; 6: 10–11. 46. Lobos G, Measey GJ. Invasive populations of Xenopus laevis (Daudin) in Chile. Herpetol J. 2002; 12: 163–168. 47. Lillo F, Marrone F, Sicilia A, Castelli G. An invasive population of Xenopus laevis (Daudin, 1802) in Italy. Herpetozoa. 2005; 18: 63–64. 48. Faraone FP, Lillo F, Giacalone G, Valvo ML. The large invasive population of Xenopus laevis in Sicily, Italy. Amphibia-Reptilia. 2008; 29: 405–412. 49. Rebelo R, Amaral P, Bernardes M, Oliveira J, Pinheiro P, LeitãoD.Xenopus laevis (Daudin, 1802), a new exotic amphibian in Portugal. Biol Invasions. 2010; 12: 3383–3387. 50. Measey GJ, Rödder D, Green SL, Kobayashi R, Lillo F, Lobos G, Rebelo R, Thirion J- M. Ongoing invasions of the African clawed frog, Xenopus laevis: a global review. Biol Invasions. 2012; 14: 2255– 2270. 51. van Sittart L, Measey GJ. Historical perspectives on global exports and research of African clawed frogs (Xenopus laevis). T Roy Soc S Afr. 2016; doi: 10.1080/0035919X.2016.1158747 52. Measey GJ, Tinsley RC. Feral Xenopus laevis in South Wales. Herpetol J. 1998; 8: 23–27. 53. Tinsley RC, Stott LC, Viney ME, Mable BK, Tinsley MC. Extinction of an introduced warm-climate alien species, Xenopus laevis, by extreme weather events. Biol Invasions. 2015; 17: 3183–3195. PMID: 26430383 54. Harrison PA, Berry PM, Butt N, New M. Modelling climate change impacts on species’ distributions at the European scale: implications for conservation policy. Environ Sci Policy. 2006; 9: 116–128.

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 16 / 19 Global Invasion Potential of Xenopus laevis

55. Randin CF, Engler R, Normand S, Zappa M, Zimmermann NE, Pearman PB, Vittoz P, Thuiller W, Gui- san A. Climate change and plant distribution: local models predict high‐elevation persistence. Glob Chang Biol. 2009; 15: 1557–1569. 56. Ihlow F, Dambach J, Engler JO, Flecks M, Hartmann T, et al. On the brink of extinction? How climate change may affect global chelonian species richness and distribution. Glob Chang Biol. 2012; 18: 1520–1530. 57. Ficetola GF, Maiorano L, Falcucci A, Dendoncker N, Boitani L, et al. Knowing the past to predict the future: land‐use change and the distribution of invasive bullfrogs. Glob Chang Biol. 2010; 16: 528– 537. 58. Kolbe JJ, Kearney M, Shine R. Modeling the consequences of thermal trait variation for the cane toad invasion of Australia. Ecol Appl. 2010; 20: 2273–2285. PMID: 21265457 59. Beaumont LJ, Hughes L, Pitman AJ. Why is the choice of future climate scenarios for species distribution modelling important? Ecol Lett. 2008; 11: 1135–146. doi: 10.1111/j.1461-0248.2008.01231.x PMID: 18713269 60. Broennimann O, Guisan A. Predicting current and future biological invasions: both native and invaded ranges matter. Ecol Lett. 2008; 4: 585–589. 61. Elith J, Leathwick JR. Species distribution models: ecological explanation and prediction across space and time. Annu Rev Ecol Syst. 2009; 40: 677–697. 62. Lobos G, Jaksic FM. The ongoing invasion of African clawed frogs (Xenopus laevis) in Chile: causes of concern. Biodivers Conserv. 2005; 14: 429–439. 63. Lillo F, Faraone FP, Lo Valvo M. Can the introduction of Xenopus laevis affect native amphibian populations? Reduction of reproductive occurrence in presence of the invasive species. Biol Invasions. 2011; 13: 1533–1541. 64. Amaral P, Rebelo R. Diet of invasive clawed frog Xenopus laevis at Lage stream (Oeiras, W Portugal). Herpetol J, 2012; 22: 187–190. 65. Lobos G, Cattan P, Estades C, Jaksic MF. Invasive African clawed frog Xenopus laevis in southern South America: key factors and predictions. Stud Neotrop Fauna Environ. 2013; 48: 1–12. 66. Measey GJ, Channing A. Phylogeography of the genus Xenopus in southern Africa. Amphibia-Repti- lia. 2003; 24: 321–330. 67. Lillo F, Dufresnes C, Faraone FP, Lo Valvo M, Stöck M. Identification and potential origin of invasive clawed frogs Xenopus (Anura: Pipidae) in Sicily based on mitochondrial and nuclear DNA. Ital J Zool. 2013; 80: 566–573. 68. Lobos G, Mendez MA, Cattan P, Jaksic F. Low genetic diversity of the successful invasive African clawed frog Xenopus laevis (Pipidae) in Chile. Stud Neotrop Fauna Environ. 2014; 49: 50–60. 69. De Busschere C, Courant J, Herrel A, Rebelo R, Rödder D, Measey GJ, Backeljau T. Unequal contribution of native South African phylogeographic lineages to the invasion of the African clawed frog, Xenopus laevis, in Europe. PeerJ. 2016; 4:e1659; doi: 10.7717/peerj.1659 PMID: 26855879 70. Furman BL, Bewick AJ, Harrison TL, Greenbaum E, Gvoždík V, Kusamba C, Evans BJ. Pan‐African phylogeography of a model organism, the African clawed frog ‘Xenopus laevis’. Mol Ecol. 2015; 24: 909–925. doi: 10.1111/mec.13076 PMID: 25583226 71. Veloz SD. Spatially autocorrelated sampling falsely inflates measures of accuracy for presence‐only niche models. J Biogeogr. 2009; 36: 2290–2299. 72. Hijmans RJ. Cross‐validation of species distribution models: removing spatial sorting bias and calibration with a null model. Ecology. 2012; 93: 679–688. PMID: 22624221 73. Boria RA, Olson LE, Goodman SM, Anderson RA. Spatial filtering to reduce sampling bias can improve the performance of ecological niche models. Ecol Model. 2014; 275: 73–77. 74. Brown JL. SDMtoolbox: a python‐based GIS toolkit for landscape genetic, biogeographic and species distribution model analyses. Methods Ecol Evol. 2014; 5: 694–700. 75. Beaumont LJ, Gallagher RV, Thuiller W, Downey PO, Leishman MR, Hughes L. Different climatic envelopes among invasive populations may lead to underestimations of current and future biological invasions. Divers Distrib. 2009; 15: 409–420. 76. Broennimann O, Guisan A. Predicting current and future biological invasions: both native and invaded ranges matter. Biol Lett. 2008; 4: 585–589. doi: 10.1098/rsbl.2008.0254 PMID: 18664415 77. Hijmans RJ, Cameron SE, Parra JL, Jones PG, Jarvis A. Very high resolution interpolated climate surfaces for global land areas. Int J Climatol. 2005; 25: 1965–1978. 78. Phillips SJ, Anderson RP, Schapire RE. Maximum entropy modeling of species geographic distributions. Ecological Modelling 190: 231–259.Pounds JA. 2001. Climate and amphibian declines. Nature. 2006;410: 639–640.

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 17 / 19 Global Invasion Potential of Xenopus laevis

79. Phillips SJ, Dudík M. Modeling of species distributions with Maxent: new extensions and a comprehensive evaluation. Ecography. 2008; 31: 161–175. 80. Fitzpatrick MC, Hargrove WW. The projection of species distribution models and the problem of non- analog climate. Biodivers Conserv. 2009; 18: 2251–2261. 81. Rocchini D, Hortal J, Lengyel S, Lengyel S, Jiménez-Valverde A, Ricotta C, Bacaro G, Chiarucci A. Accounting for uncertainty when mapping species distributions: the need for maps of ignorance. Prog Phys Geog. 2011; 35: 211–226. 82. Swets JA. Measuring the accuracy of diagnostic systems. Science. 1988; 240: 1285–1293. PMID: 3287615 83. VanDerWal J, Shoo LP, Graham C, Williams SE. Selecting pseudo-absence data for presence-only distribution modeling: how far should you stray from what you know? Ecol Model. 2009; 220: 589–594. 84. Thuiller W, Münkemüller T, Lavergne S, Mouillot D, Mouquet N, Schiffers K, Gravel D. A road map for integrating eco‐evolutionary processes into biodiversity models. Ecol Lett. 2013; 16: 94–105. doi: 10. 1111/ele.12104 PMID: 23679011 85. R Development Core Team. R: A language and environment for statistical computing. Available: http://www.Rproject.org (15 February 2013). 86. Allouche O, Tsoar A, Kadmon R. Assessing the accuracy of species distribution models: prevalence, kappa and the true skill statistic (TSS). J Appl Ecol. 2006; 43: 1223–1232. 87. Pearson RG, Raxworthy CJ, Nakamura M, Peterson AT. Predicting species distributions from small numbers of occurrence records: a test case using cryptic geckos in Madagascar. J Biogeogr. 2007; 34: 102–117. 88. Liu C, Berry PM, Dawson TP, Pearson RG. Selecting thresholds of occurrence in the prediction of species distributions. Ecography. 2005; 28: 385–393. 89. Webber BL, Yates CJ, Le Maitre DC, Scott JK, Kriticos DJ, Ota N, Mc Neil A, Le Roux JK, Midgley GF. Modelling horses for novel climate courses: insights from projecting potential distributions of native and alien Australian acacias with correlative and mechanistic models. Divers Distrib. 2011; 17: 978– 1000. 90. Elith J, Kearney M, Phillips S. The art of modelling range‐shifting species. Meth Ecol Evol. 2010; 1: 330–342. 91. Moss RH, Edmonds JA, Hibbard KA, Manning MR, Rose SK, et al. The next generation of scenarios for climate change research and assessment. Nature. 2010; 463: 747–756. doi: 10.1038/nature08823 PMID: 20148028 92. Peterson AT, Nyari AS. Ecological niche conservatism and Pleistocene refugia in the Thrush-like Mourner, Schiffornis sp., in the Neotropics. Evolution. 2008; 62: 173–183. PMID: 18005155 93. Ramirez-Villegas J, Jarvis A. Downscaling global circulation model outputs: the delta method decision and policy analysis Working Paper No. 1. J Pol Anal Manag. 2010; 1: 1–18. 94. Bertelsmeier C, Luque GM, Hoffmann BD, Courchamp F. Worldwide ant invasions under climate change. Biodivers Conserv. 2014; 24: 117–128. 95. Balinsky B. I. (1969). The reproductive ecology of amphibians of the Transvaal highveld. Zoologica africana, 4(1), 37–93. 96. McCoid MJ. An observation of reproductive-behaviour in a wild population of African clawed frogs, Xenopus laevis, in California, Calif fish game. 1985; 71: 245–246. 97. McCoid MJ, Fritts TH. Observations of feral populations of Xenopus laevis (Pipidae) in southern Cali- fornia. Bull—South Calif Acad Sci. 1980; 79: 82–86. 98. Herrel A, van der Meijden A. An analysis of the live reptile and amphibian trade in the USA compared to the global trade in endangered species. Herpetol J. 2014; 24: 103–110. 99. Meyerson LA, Reaser JK. Biosecurity: moving toward a comprehensive approach. Bioscience. 2002; 52: 593–600. 100. Vogel G. Proposed frog ban makes a splash. Science. 2008; 319: 1472–1472. 101. Fouquet A, Measey GJ. Plotting the course of an African clawed frog invasion in Western France. Anim Biol. 2006; 56: 95–102. 102. De Villiers, FA. The dispersal ability, performance and population dynamics of Cape Xenopus frogs. MSc Thesis, Stellenbosch University, 2016. 103. Crayon JJ. Species account: Xenopus laevis. In: Lannoo MJ, editor. Amphibian declines: the conservation status of United States species. University of California Press, Berkeley; 2005. Pp. 522–25. 104. McNeely JA, Mooney HA, Neville LE, Schei P, Waage JK, editors. Global strategy on invasive alien species. Gland, Switzerland: IUCN on behalf of the Global Invasive Species Programme. 2001.

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 18 / 19 Global Invasion Potential of Xenopus laevis

105. National Invasive Species Council. Meeting the invasive species challenge: national invasive species management plan. Washington, DC: National Invasive Species Council. 2001.

PLOS ONE | DOI:10.1371/journal.pone.0154869 June 1, 2016 19 / 19

APPENDIX C

Courant J, Vogt S, Marques R, et al (2017) Are invasive populations characterized by a broader diet than native populations? PeerJ 5:e3250.

Are invasive populations characterized by a broader diet than native populations?

Julien Courant1, Solveig Vogt2,3, Raquel Marques4, John Measey3, Jean Secondi5,6, Rui Rebelo4, André De Villiers3, Flora Ihlow2, Charlotte De Busschere7, Thierry Backeljau7,8, Dennis Rödder2 and Anthony Herrel1,9

1 UMR 7179, Département d’Ecologie et de Gestion de la Biodiversité, Centre National de la Recherche Scientifique, Paris, France 2 Herpetology Section, Zoologisches Forschungsmuseum Alexander Koenig, Bonn, Germany 3 Centre for Invasion Biology, Stellenbosch University, Stellenbosch, South Africa 4 Centre for Ecology, Evolution and Environmental Changes, Faculdade de Ciências da Universidade de Lisboa, Lisboa, Portugal 5 UMR5023 Ecologie des Hydrosystèmes Naturels et Anthropisés, ENTPE, CNRS, Université de Lyon, Université Lyon 1, Villeurbanne, France 6 UMR 6554 LETG –LEESA, Université d’Angers, Angers, France 7 Royal Belgian Institute of Natural Sciences, Brussels, Belgium 8 Evolutionary Ecology Group, University of Antwerp, Antwerp, Belgium 9 Evolutionary Morphology of Vertebrates, Ghent University, Ghent, Belgium

ABSTRACT Background. Invasive species are among the most significant threats to biodiversity. The diet of invasive animal populations is a crucial factor that must be considered in the context of biological invasions. A broad dietary spectrum is a frequently cited characteristic of invasive species, allowing them to thrive in a wide range of environments. Therefore, empirical studies comparing diet in invasive and native populations are necessary to understand dietary requirements, dietary flexibility, and the associated impacts of invasive species. Methods. In this study, we compared the diet of populations of the African clawed frog, Xenopus laevis in its native range, with several areas where it has become invasive. Submitted 18 January 2017 Each prey category detected in stomach contents was assigned to an ecological category, Accepted 30 March 2017 Published 16 May 2017 allowing a comparison of the diversity of ecological traits among the prey items in the diet of native and introduced populations. The comparison of diets was also performed Corresponding author Julien Courant, using evenness as a niche breadth index on all sampled populations, and electivity as a [email protected], prey selection index for three out of the six sampled populations. [email protected] Results. Our results showed that diet breadth could be either narrow or broad in Academic editor invasive populations. According to diet and prey availability, zooplankton was strongly David Roberts preferred in most cases. In lotic environments, zooplankton was replaced by benthic Additional Information and preys, such as ephemeropteran larvae. Declarations can be found on Discussion. The relative proportions of prey with different ecological traits, and page 12 dietary variability within and between areas of occurrence, suggest that X. laevis is DOI 10.7717/peerj.3250 a generalist predator in both native and invasive populations. Shifts in the realized trophic niche are observed, and appear related to resource availability. Xenopus laevis Copyright 2017 Courant et al. may strongly impact aquatic ecosystems because of its near complete aquatic lifestyle and its significant consumption of key taxa for the trophic relationships in ponds. Distributed under Creative Commons CC-BY 4.0 Subjects Biodiversity, Ecology, Ecosystem Science, Zoology OPEN ACCESS Keywords Diet breadth, Trophic niche, African clawed frog, Invasive, Electivity, Native

How to cite this article Courant et al. (2017), Are invasive populations characterized by a broader diet than native populations? PeerJ 5:e3250; DOI 10.7717/peerj.3250 INTRODUCTION Invasive species usually occupy a wide geographical range in their native area. Invasive species are typically characterized by a number of traits that favor the establishment and spread across new ecosystems, including a broad environmental tolerances, high genetic variability, rapid growth, early sexual maturity combined with a high reproductive rate, short generation time, broad diet, gregariousness, rapid dispersal, and they are often commensal (Ricciardi & Rasmussen, 1998). Of course, not all invasive species meet all these criteria (Lodge, 1993). For example, successful invaders do not necessarily exhibit a broad diet (Vazquez, 2006). Yet a large dietary niche breadth is frequently considered as a hallmark of an invasive taxon. The dietary niche is a component of the Eltonian niche, defined as the position of an organism, exhibiting a null population growth rate, in the trophic relationships with others organisms of the ecosystem such as its nutrients, predators and competitors (Chase & Leibold, 2003). Another aspect of the ecological niche is the Grinnellian niche, defined as the set of all values of the abiotic parameters enabling the occupancy of an area by a species (Soberón, 2007). The Grinnellian niche has known a recent intensification of research with the development of species distribution models (e.g., Angetter, Lötters & Rödder, 2011; Guisan et al., 2013). Based on a set of occurrence records and predictor variables, these models determine a given species fundamental niche and facilitate the assessment of potential niche shifts when projected onto novel conditions. While these models are based on the assumption that species retain their ancestral traits over time (see Ackerly, 2003; Wiens & Graham, 2005) recent evidence (Broennimann et al., 2007; Mukherjee et al., 2012; Stiels et al., 2015) revealing shifts in realized Grinnellian niches on a macro-ecological scale call this concept into question. How far Eltonian niches operating on a population level are variable, under the assumption of niche conservatism, is less well studied. Whether a population maintains its characteristics, or shifts them in the course of the invasion process, will contribute determining its ecological impact. This information is therefore of crucial importance for conservation practitioners facing the threat posed by invasive species. Native to southern Africa, the African clawed frog, Xenopus laevis, has been introduced in many countries on four continents, where accidentally and deliberately released individuals have established viable populations (see Measey et al., 2012). Despite its importance as a biological model organism (Cannatella & De Sa, 1993), and the abundance of invasive populations, few field studies have been undertaken in colonized ranges (for a review, see Measey et al., 2012). Xenopus laevis has been reported to negatively affect the invaded ecosystems, and as a consequence has been ranked as having the second greatest impact on native ecosystems by any amphibian (Measey et al., 2016). Shifts in the Grinnellian niche of X. laevis have been recently demonstrated (Rödder et al., in press), whereas studies on changes in the Eltonian niches have not yet been undertaken. The diet of X. laevis has been studied in the species’ native range of South-Africa (Schoonbee, Prinsloo & Nxiweni, 1992), as well as in several introduced populations in the United States of America (USA) (McCoid & Fritts, 1980), Wales (Measey, 1998a), Chile (Lobos & Measey, 2002), Italy (Faraone et al., 2008), Portugal (Amaral & Rebelo, 2012), and

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 2/16 France (Courant et al., 2014). In most studies, the majority of prey items are aquatic, with zooplankton and dipteran larvae being the most frequent. Measey (1998b) also noticed the importance of terrestrial prey. While stomach content analyses conducted in Portugal (Amaral & Rebelo, 2012) and in the USA (McCoid & Fritts, 1980) revealed that X. laevis consumed eggs of fishes and amphibians, no study has reported a direct impact linked to predation. A similar study, conducted in South African aquaculture ponds, revealed that farmed fish larvae constituted a large proportion (5–25% occurrence frequency according to fish size) of frog stomach contents (Schramm, 1987), while another study found X. laevis to consume large quantities of anuran eggs and larvae (Vogt et al., 2017). Given the wide diversity of prey items, dietary studies on X. laevis usually suggest a generalist feeding behaviour, but only one study has thus far investigated prey electivity (Measey, 1998a). No previous studies have explicitly compared the feeding behavior of populations in different ecological contexts, and with different invasion histories. In this study, we compiled published data on the diet of X. laevis from the USA (McCoid & Fritts, 1980), Wales (Measey, 1998a), Chile (Lobos & Measey, 2002), and South Africa (Vogt et al., 2017) with data collected during recent field work in Portugal and France to test the hypotheses that (i) trophic niche breadth is wider in native populations, hence releaving the capacity of the species to readily adapt to novel environments; and (ii) the diet of invasive populations differs significantly between the invaded ranges depending on local prey availability, and thus resulting in a low degree of electivity and population-specific niche shifts.

MATERIAL & METHODS Data sampling Our dataset comprised 1,458 individuals from six countries, across four continents (Table 1). In most areas (Chile, South Africa, Wales, France), frogs were caught using funnel traps. In Portugal, animals were captured using electrofishing because the colonized habitats, mainly fast flowing streams, prevented the use of traps. In South Africa, research permission was issued by CapeNature (AAA007-01867) and SANParks (RC/2014-2015/001–2009/V1), with ethics clearance from Stellenbosch University Research Ethics Committee: Animal Care & Use (SU-ACUD15-00011). Animals from the Portuguese invasive population were captured under permit no 570/2014/CAPT from Instituto da Conservacão¸ da Natureza e das Florestas, in the scope of the ‘‘Plano de erradicacão¸ de Xenopus laevis nas ribeiras do Concelho de Oeiras’’ (Eradication plan of Xenopus laevis in the streams of Oeiras Municipality). In France, a research permit was provided by the prefecture of the Deux- Sèvres department. Stomach content samples were obtained either by stomach flushing or dissection, following euthanasia of individuals by lethal injection of sodium pentobarbital or immersion in MS222. We considered that analyzing and comparing data collected with both dissection and flushing methods were valid and did not induce any bias (Wu, Li & Wang, 2007).

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 3/16 Table 1 Characteristics of the methods used to capture and describe the diet of Xenopus laevis.

South Africa Wales France Chile Portugal USA Population status Native Extincta Invasive Invasive Invasive Invasive Period of capture From 06/2014 From 05/1995 From 05/2014 01/1998 From 06/2014 1975–1976 To-09/2014 To 08/1996 To 10/2014 03/2001 To 08/2014 Geographical coordinates Latitude/Longitude NWb S34◦18′24′′ N51◦27′33′′ N47◦16′14′′ S33◦29′ N38◦45′09′′ See McCoid & E18◦25′35′′ W3◦33′11′′ W0◦33′56′′ W70◦54′ W9◦17′27′′ Fritts (1980) Latitude/Longitude SEb S34◦20′06′′ NA N46◦53′41′′ S33◦37′ N38◦42′35′′ See McCoid & ′′ ′′ E19◦04′29 W0◦31′11 W70◦39′ W9◦16′25′′ Fritts (1980) Sampling design Method Trap Trap Trap Trap Electrofishing NA Capture occasion/site From 1 to 4 29 3 1 From 1 to 4 1 Number of sites 8 1 26 2 12 1 Number of individuals 164 375 438 48 352 81 Prey availability Yes Yes Yes No No No Habitat type Ponds Pond Ponds Ponds Streams Streams Prey collection method Flushing/ Flushing Dissection Dissection Dissection Dissection Dissection Published data Prey frequency in stomachs Yes Yes No Yes No Yes Niche breadth Yes No No No No No Electivity Yes Yes No No No No Individual data Yes No Yes Yes Yes No

Notes. aThe population introduced in Wales went extinct twenty years after the data collection used in our study (Tinsley et al., 2015). b Geographical coordinates (WGS 84), northwestern (NW) and southeastern corners (SE), of the minimum rectangle encompassing all sampled sites for Ns > 1. Data analysis Prey items retrieved from the stomach content samples were identified to the lowest taxonomic level possible. However, for analytical consistency, we retained the lowest common taxonomic level that could be identified for all prey items. As volume and mass of prey items were not available for most studies, analyses were performed using prey frequencies. Each taxonomic prey category was assigned to one of the following ecological traits: plankton, benthos, nekton and terrestrial. Some groups of invertebrates belong to different ecological trait categories depending on their life stage, e.g., aquatic in their larval stage and terrestrial in their adult stage (Diptera, Ephemeroptera, Trichoptera). Thus, adults and larvae were treated separately when assigned to different ecological traits even though they belong to the same taxonomic prey category. The diet of populations was first compared by calculating the relative abundance of each prey category and ecological trait. The relative abundance of prey classes (aquatic invertebrates, terrestrial invertebrates, vertebrates) was also calculated. To assess variation in diet between populations, a Principal Component Analysis (PCA) was performed on reduced and centered relative abundances of each prey category.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 4/16 The concept of niche breadth can be applied to comparative studies of diet between populations or species (Slatyer, Hirst & Sexton, 2013), even if it has not always been treated within this contextual vocabulary (e.g., Rehage, Barnett & Sih, 2005; Luiselli et al., 2007; Dalpadado & Mowbray, 2013). We calculated niche breadth for all populations using the evenness measure J’. This index is based on the Shannon–Wiener’s index H’(Shannon & Weaver, 1964), as recommended by Colwell & Futuyma (1971):

−Ppi ∗log pi J ′ = log n

where pi is the proportion of the prey category i in the diet and n is the number of food categories. To test whether J ′ is affected by the number of study sites the relationship of both variables was assessed using a nonlinear regression. Prey availability was quantified in habitats for three of the six populations (France, Wales and South Africa) included in this study. Following Measey (1998a), the same sampling method was applied to all populations. Prey selection was assessed using the Vanderploeg & Scavia’s (1979) relativized electivity index recommended by Lechowicz (1982):

1 ri Wi − p E∗ = n with W = i + 1 i r /p Wi n P i i

where ri is the relative abundance of prey category i in the diet and pi is the relative abundance of prey category i in the environment. The number of prey categories included in the analysis is represented by n.

RESULTS Across all samples zooplankton was the most common prey type with a mean relative abundance of 56.21% (Standard Deviation = 32.80%), followed by ephemeropteran larvae (10.31% ± 23.40%), dipteran larvae (9.68% ± 7.23%), and gastropods (7.24% ± 6.07%). The fifth and sixth most represented prey items were amphibian eggs, excluding X. laevis (4.86% ± 11.08%) and X. laevis eggs (3.46% ± 7.55%) respectively. Three aquatic invertebrate orders (Coleoptera, Diptera, and Heteroptera) were detected in all study sites (Table 2), while most of the terrestrial categories were exclusively found in one or two sites. Cannibalism of larvae and/or eggs was recorded in every locality, except Chile. Aquatic invertebrates represented the most consumed prey item class, with a relative abundance ranging from 66% in South Africa to 99% in Wales. Terrestrial invertebrates were rarely consumed and consequently relative abundance ranged from 0.02% in France and the USA to 1.5% in Chile. Variability in relative abundance was highest for vertebrate prey, reaching a maximal relative abundance of 33% in South Africa while being absent in Chile. According to the PCA (Fig. 1), the diet of the Portuguese and South African populations was respectively characterized by high relative abundances of ephemeropteran larvae, and eggs of native amphibians. The cluster representing Wales and Chile was characterized by a high occurrence of zooplankton and a very low occurrence of all vertebrate prey

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 5/16 Table 2 Relative abundance (in %) of the prey items identified in the native and colonized ranges of Xenopus laevis. When prey items are observed in very low quantities (N < 3), they are noted as <0.01% in the table. The main prey categories of each populations are underlined. Below the name of each ecological category, we mention the mean and the standard deviation of the relative abundance of items found in stomach contents.

Cat no South Africa Wales France Chile Portugal USA Aquatic invertebrates 65.99 99.15 80.68 98.54 96.56 98.13 1 Zooplankton 51.87 92.55 39.62 82.51 2.28 68.89 Benthos 8.86 6.45 37.08 14.85 93.90 27.41 34.80% ± 34.17% 2 Annelida 0.03 0.00 0.02 0.00 0.05 0.00 3 Turbellaria 0.00 <0.01 0.00 0.00 0.00 0.00 4 Gastropoda 0.00 0.00 9.82 10.64 15.05 7.99 5 Bivalvia 0.00 0.24 0.44 0.00 0.05 0.00 6 Acari 4.63 0.00 0.00 0.00 0.05 0.00 8 Amphipoda 1.83 0.02 3.09 0.00 0.05 4.21 9 Isopoda 0.00 0.00 0.00 0.00 2.62 0.05 10 Decapoda 0.00 0.00 0.00 0.00 0.06 0.00 11 Diptera (larvae) 1.78 6.14 20.90 4.21 15.24 9.91 12 Ephemeroptera (larvae) 0.00 0.05 2.79 0.00 58.02 0.98 13 Trichoptera (larvae) 0.59 0.00 0.02 0.00 2.71 4.27 Nekton 5.26 0.14 4.00 1.19 0.50 1.83 7.74% ± 13.87% 14 Coleoptera (larvae) 1.97 0.03 0.02 0.00 0.06 0.27 15 Coleoptera (adult) 0.67 0.09 1.74 0.44 0.38 0.34 16 Heteroptera 1.09 0.02 1.12 0.18 0.01 0.16 17 Zygoptera (larvae) 0.79 0.00 0.38 0.44 0.00 0.53 18 Anisoptera (larvae) 0.74 0.00 0.74 0.13 0.05 0.53 Terrestrial invertebrates 0.86 0.43 0.23 1.46 0.70 0.24 0.64% ± 0.46% 19 Arachnida 0.46 0.01 0.00 1.21 0.05 0.11 20 Isopoda 0.00 0.05 0.00 0.03 0.04 0.00 21 Chilopoda 0.00 0.00 0.00 0.00 0.01 0.00 22 Diplopoda 0.00 0.00 0.00 0.00 0.01 0.00 23 Diptera 0.09 0.06 0.00 0.00 0.04 0.08 24 Neuroptera 0.04 0.00 0.00 0.00 0.00 0.00 25 Hymenoptera 0.27 0.00 0.00 0.05 0.16 0.00 26 Coleoptera 0.00 0.01 0.05 0.00 0.00 0.05 27 Lepidoptera (larvae) 0.00 <0.01 0.00 0.00 0.00 0.00 28 Lepidoptera (adult) 0.00 0.00 0.00 0.15 0.00 0.00 29 Dermaptera 0.00 0.01 0.00 0.03 0.00 0.00 30 Heteroptera 0.00 0.00 0.02 0.00 0.10 0.00 31 Annelida 0.00 0.02 0.09 0.00 0.02 0.00 32 Orthoptera 0.00 0.00 0.08 0.00 0.00 0.00 33 Aphids 0.00 0.27 0.00 0.00 0.00 0.00 (continued on next page)

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 6/16 Table 2 (continued) Cat no South Africa Wales France Chile Portugal USA 34 Trichoptera 0.00 0.00 0.00 0.00 0.02 0.00 35 Ephemeroptera 0.00 0.00 0.00 0.00 0.24 0.00 Vertebrates 33.13 0.41 19.09 0.00 2.74 1.63 9.45% ± 13.54% 36 Fish (adult) 0.00 0.00 0.11 0.00 0.02 0.08 37 Fish (egg) 0.00 0.00 0.00 0.00 0.82 0.00 38 Amphibia (adult) 0.14 0.00 0.02 0.00 0.00 0.00 3 Amphibia (larvae) 3.37 0.00 0.00 0.00 0.00 0.00 39 Amphibia (egg) 27.62 0.00 0.08 0.00 1.65 0.00 40 X. laevis (larvae) 0.04 0.00 0.08 0.00 0.01 0.03 41 X. laevis (egg) 0.00 0.41 18.81 0.00 0.00 1.52 42 Amphibia (rest) 1.98 0.01 0.00 0.00 0.22 0.00 43 Bird (feather) 0.00 <0.01 0.00 0.00 <0.01 0.00 44 Mammals 0.00 <0.01 0.00 0.00 0.00 0.00

Portugal

France

USA

Wales 0

Chile Dim2 (23%)

−1

South Africa

−2

−2 −1 0 1 2 Dim1 (24.7%)

Figure 1 Principal components of the diet of the native (South Africa) and invasive populations of Xenopus laevis, with prey categories as individuals (dots, squares, triangles and crosses) and populations as variables (black arrows).

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 7/16 Relative Abundance (%) Abundance Relative 0 20 40 60 80 100

South Africa Wales France Chile Portugal USA

Figure 2 Occurrence of the main ecological traits among the native population (South Africa) and the five populations of Xenopus laevis. Terrestrial prey (black), benthic prey (dark blue), nektonic items (sky blue) and planktonic prey (cyan).

items. Except for the eggs of X. laevis, French and American samples share the three most abundant prey categories (i.e., Diptera, Gastropoda, and Zooplankton), which explains the short distance between these samples in the PC space. Benthic taxa represented 8.54% of the prey categories in South Africa, 6.87% in Wales, 14.85% in Chile, 28.8% in the USA, 55.87% in France, and 93.89% in Portugal (Fig. 2). Nektonic taxa represented 39.14% of the prey items in South Africa, 0.14% in Wales, 1.18% in Chile, 1.94% in the USA, 4.20% in France, and 3.01% in Portugal. Standardized evenness was 0.48 in South Africa, 0.10 in Wales, 0.27 in Chile, 0.42 in the USA, 0.55 in France, 0.41 in Portugal (Fig. 3A). The relationship between ′ evenness (J ) and the number of sampled sites (Ns) followed a logarithmic function ′ (J = 0.086 ∗ ln(Ns) + 0.249; Fig. 3B). The slope of the curve decreased from one to ten sites, and stabilized at 25 sites. There was no correlation between trophic niche breadth estimations and the number of stomach contents analyzed per population. Negative electivity values were observed for most prey items (Fig. 4). In contrast, zooplankton was preferred (positive electivity) in the Welsh, South African and French populations. The electivity values of the other prey categories were variable, indicating that they were either selected, avoided, or not represented in the sampling.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 8/16 J’ J’ (A) 0,6 (B)

0,5

0,4

0,3

0,2

0,1

0.0 0.1 0.2 0.3 0.4 0.5 0.6 0 NS South Africa Wales France Chile Portugal USA 0 5 10 15 20 25 30

Figure 3 Niche breadth calculated for diet data in native and colonized ranges of Xenopus laevis. Cal- ′ ′ culated with the Evenness J (A) and relationship between J and the number of sites NS used in localities (B).

DISCUSSION In studies focusing on invasive species, both niche conservatism and niche shifts are commonly reported in the literature (Tillberg et al., 2007; Caut, Angulo & Courchamp, 2008; Comte, Cucherousset & Olden, 2016). Here, we report strong modifications in the realized dietary niche between naturally occurring and introduced or invasive populations of X. laevis. While these differences mainly constituted contractions or expansions of the realized dietary niche, the diet of the Portuguese population represented a shift in the species’ diet. Individuals from this population were captured in fast flowing streams using electrofishing, while all others frogs were caught in lentic environments. The difference might therefore be attributed to the habitat characteristics, reinforcing the hypothesis that X. laevis is a generalist predator that modifies its diet according to available resources. Our study provides the first analysis of prey availability for a large number of sites that include both natural and invaded areas. Our findings indicate that X. laevis may expand or shift some dimensions of its trophic niche in novel environments. This result is significant, and has clear consequences for evaluating conditions that favor the invasion of newly introduced populations. Suitable trophic conditions allowing a positive population growth rate may be found in many places where prey abundance is sufficient. A recent macroecological assessment revealed that large areas of suitable climatic conditions were available outside the species’ native range, and that X. laevis was likely to expand its range in Europe as a result of climate change (Ihlow et al., 2016). The broad global trophic niche of X. laevis and its ability to adapt its diet according to local conditions, contribute to the strong invasive potential of this species, and the high impact it may induce on its environment (see Kumschick et al., 2017). The diet of X. laevis has been studied in its native range (Schoonbee, Prinsloo & Nxiweni, 1992; Vogt et al., 2017), and in different invasive populations (McCoid & Fritts, 1980;

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 9/16 Electivity -1.0-1,0 -0,8-0.8 -0.6 -0,6-0.4 -0,4 -0,2-0.2 0.0 0,0 0,20.2 0.4 0,4 0,60.6 0,80.8 1.0 1,0

Zooplankton

Gastropoda

Acari

Amphipoda

Coleoptera

Diptera

Heteroptera

Zygoptera

Anisoptera

Ephemeroptera

Trichoptera

Amphibia

Oligochaeta

Figure 4 Electivity index for each aquatic prey category consumed in the native population of Xenopus laevis in South Africa (brown) and the invasive populations of Wales (dark-orange) and France (light- orange).

Measey, 1998a; Lobos & Measey, 2002; Faraone et al., 2008; Amaral & Rebelo, 2012; Courant et al., 2014). The first study carried out in the native range was performed in a fish farm (Schoonbee, Prinsloo & Nxiweni, 1992), which does not necessarily represent the typical diet of native populations. In most studies, including ours, X. laevis was found to predominantly

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 10/16 consume relatively small prey items. In Portugal, X. laevis mostly inhabits streams where zooplankton is rare and the main prey items are benthic ephemeropteran larvae. Prey availability in streams was not studied in Portugal, but collecting such data may help understanding the dietary shift we observed in this population. In all populations but Chile, predation on amphibians was observed, including X. laevis itself, which represented the most frequent vertebrate taxon in the collective sample. This cannibalistic behavior has been recorded in non-native populations, including in the USA (McCoid & Fritts, 1980), Wales (Measey, 1998a), Chile (Lobos & Jaksic, 2005), and Italy (Faraone et al., 2008) but the high frequencies observed in France and South Africa are unprecedented. In the Chilean population, autochthonous amphibians, as well as X. laevis eggs and larvae, were not observed during the sampling period (J Measey, pers. comm., 2016) which may explain their absence in the diet. The small number of stomach contents sampled and sites analyzed for this population may also provide a biased assessment of the putative absence of native amphibians in this area. Predation on amphibians was minor in most populations, except for the native range of X. laevis. This low occurrence of predation on amphibians in other localities may be related to the season during which studies were carried out, and possible changes in the behavior of native amphibians which have co-existed with X. laevis for decades. Our study does not provide any evidence supporting this idea, but ongoing studies should bring new insights into this question. The noteworthy anurophagy reported in this study corroborates the conclusions of Measey et al. (2016) regarding the occurrence of this behavior among pipids. In this study, terrestrial invertebrates represented the least consumed prey class. However, cumulatively, there were as many taxonomic categories among terrestrial, as among aquatic invertebrates. The occurrence of terrestrial prey items did not vary much between populations, suggesting that there were no local specializations for the capture of terrestrial prey. In previous dietary studies of X. laevis, authors concluded that the high portion of terrestrial prey could not exclusively be explained by the capture of terrestrial invertebrates that had fallen into the water (Measey, 1998b). From a methodological perspective, sampling effort in each locality was heterogeneous with respect to the sampling period, the number of prospected sites (range: 1–26), and the number of individuals analyzed per population. Consequently, niche breadth could be under-estimated in populations where only few sites were sampled (Wales, Chile, and USA) or where sites were inter-connected (streams in Portugal). According to our results, there is a positive relationship between diet breadth and the number of prospected sites. As no threshold was identified for an optimal number of study sites, we would recommend using as many spatial and temporal replicates as possible for studies aiming at comparing niche breadths. The diet of an individual may be influenced by its size (Schafer et al., 2002), age (Gales, 1982; Rutz, Whittingham & Newton, 2006), and sex independently of size (Gales, 1982; Gö¸cmen et al., 2011; Van Ngo, Lee & Ngo, 2014). In our study, these data were not available for every population, preventing us from analyzing their effect on diet breadth. In other respects, the unique native population included in our study may not be representative of the diet in the native range. Estimations of niche breadth and prey selectivity may have differed had they been based on a larger sample of native populations.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 11/16 Therefore, we encourage future investigators to consider as many naturally occurring populations as invasive populations in studies aimed at understanding the feeding ecology of invasive animals. Our results indicate that no prey categories are strongly selected, except for zooplankton in Wales. This suggests that X. laevis does not usually specializes its diet and hence does not develop a population specific dietary niche. This characteristic may enhance its capacity to establish and spread in novel environments. Potential perturbations of X. laevis in its environment, linked to the large predation on small prey items, are crucial elements of the trophic relationships in aquatic ecosystems that still need to be demonstrated. Our results reflect the diet of X. laevis in invasive populations after decades of colonization and do not necessarily reflect the diet of the species at the moment of introduction. Some invasive species modify their diet during the years, or decades following habitat colonization (e.g., Tillberg et al., 2007; Gkenas et al., 2016). Comparing the diet of individuals at the core and the edge of a newly colonized area may be an effective approach to investigate the change in dietary composition of X. laevis during the colonization process.

ACKNOWLEDGEMENTS We are sincerely grateful to the two anonymous reviewers and the Academic Editor for their helpful and constructive comments on the manuscript.

ADDITIONAL INFORMATION AND DECLARATIONS

Funding This research was funded by BiodivERsA BR/132/A1/INVAXEN-BE ‘‘Invasive biology of Xenopus laevis in Europe: ecology, impact and predictive models’’, the DST-NRF Centre of Excellence for Invasion Biology and the National Research Foundation of South Africa (NRF Grant No. 87759) to John Measey. The funders had no role in study design, data collection and analysis, decision to publish, or preparation of the manuscript. Grant Disclosures The following grant information was disclosed by the authors: DST-NRF Centre of Excellence for Invasion Biology and National Research Foundation of South Africa: 87759. Competing Interests John Measey is an Academic Editor for PeerJ. Author Contributions • Julien Courant conceived and designed the experiments, performed the experiments, analyzed the data, contributed reagents/materials/analysis tools, wrote the paper, prepared figures and/or tables, reviewed drafts of the paper. • Solveig Vogt, Raquel Marques and André De Villiers performed the experiments, wrote the paper, reviewed drafts of the paper.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 12/16 • John Measey conceived and designed the experiments, performed the experiments, contributed reagents/materials/analysis tools, wrote the paper, reviewed drafts of the paper. • Jean Secondi and Anthony Herrel conceived and designed the experiments, contributed reagents/materials/analysis tools, wrote the paper, reviewed drafts of the paper. • Rui Rebelo conceived and designed the experiments, performed the experiments, wrote the paper, reviewed drafts of the paper. • Flora Ihlow, Charlotte De Busschere, Thierry Backeljau and Dennis Rödder wrote the paper, reviewed drafts of the paper. Animal Ethics The following information was supplied relating to ethical approvals (i.e., approving body and any reference numbers): In South Africa, research permission was issued by CapeNature (AAA007-01867) and SANParks, with ethics clearance from Stellenbosch University Research Ethics Committee: Animal Care & Use (SU-ACUD15-00011). Animals from the Portuguese invasive population were captured under permit no 570/2014/CAPT from Instituto da Conservacão¸ da Natureza e das Florestas, in the scope of the ‘‘Plano de erradicacão¸ de Xenopus laevis nas ribeiras do Concelho de Oeiras ("Eradication plan of Xenopus laevis in the streams of Oeiras Municipality").’’ In France, a research permit was provided by the prefecture of the Deux-Sèvres department. Data Availability The following information was supplied regarding data availability: The raw data has been supplied as a Supplementary File. Supplemental Information Supplemental information for this article can be found online at http://dx.doi.org/10.7717/ peerj.3250#supplemental-information.

REFERENCES Ackerly DD. 2003. Community assembly, niche conservatism, and adaptive evolution in changing environments. International Journal of Plant Sciences 164:165–184. Amaral P, Rebelo R. 2012. Diet of invasive clawed frog Xenopus laevis at Lage stream (Oeiras, W Portugal). Herpetological Journal 22:187–190. Angetter LS, Lötters S, Rödder D. 2011. Climate niche shift in invasive species: the case of the brown anole. Biological Journal of the Linnean Society 104:943–954 DOI 10.1111/j.1095-8312.2011.01780.x. Broennimann O, Treier UA, Müller-Schärer H, Thuiller W, Peterson AT, Guisan A. 2007. Evidence of climatic niche shift during biological invasion. Ecology Letters 10:701–709 DOI 10.1111/j.1461-0248.2007.01060.x. Cannatella DC, De Sa RO. 1993. Xenopus laevis as a model organism. Systematic Biology 42:426–507 DOI 10.1093/sysbio/42.4.476.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 13/16 Caut S, Angulo E, Courchamp F. 2008. Dietary shift of an invasive predator: rats, seabirds and sea turtles. Journal of Applied Ecology 45:428–437 DOI 10.1111/j.1365-2664.2007.01438.x. Chase JM, Leibold MA. 2003. Ecological niches: linking classical and contemporary approaches. Chicago: The University of Chicago Press. Colwell RK, Futuyma DJ. 1971. On the measurement of niche breadth and overlap. Ecology 52:567–576 DOI 10.2307/1934144. Comte L, Cucherousset J, Olden JD. 2016. Global test of Eltonian niche conservatism of nonnative freshwater fish species between their native and introduced ranges. Ecography 39:1–9 DOI 10.1111/ecog.02007. Courant J, Thirion J, Guillon M, Grillet P, Grosselet O. 2014. Le régime alimentaire de Xenopus laevis (Daudin, 1802) (Anura: Pipidae) introduit en France. Bulletin De La Société Herpétologique De France 150:1–7. Dalpadado P, Mowbray F. 2013. Comparative analysis of feeding ecology of capelin from two shelf ecosystems, off Newfoundland and in the Barents Sea. Progress in Oceanography 114:97–105 DOI 10.1016/j.pocean.2013.05.007. Faraone FP, Lillo F, Giacalone G, Lo Valvo M. 2008. The large invasive population of Xenopus laevis in Sicily, Italy. Amphibia-Reptilia 29:405–412 DOI 10.1163/156853808785112075. Gales RP. 1982. Age- and sex-related differences in diet selection by Rattus rattus on Stewart Island, New Zealand. New Zealand Journal of Zoology 9:463–466 DOI 10.1080/03014223.1982.10423878. Gkenas C, Magalhães MF, Cucherousset J, Domingos I, Ribeiro F. 2016. Long term patterns in the late summer trophic niche of the invasive pumpkinseed sunfish Lepomis gibbosus. Knowledge and Management of Aquatic Ecosystems 417:0–5 DOI 10.1051/kmae/2016006. Göcmen¸ B, Cicek K, Yildiz MZ, Atatur MK, Dincaslan YE, Mebert K. 2011. A preliminary study on the feeding biology of the dice snake, Natrix tessellata, in Turkey. Mertensiella 18:365–369. Guisan A, Tingley R, Baumgartner JB, Naujokaitis-Lewis I, Sutcliffe PR, Tulloch AIT, Regan TJ, Brotons L, Mcdonald-Madden E, Mantyka-Pringle C, Martin TG, Rhodes JR, Maggini R, Setterfield SA, Elith J, Schwartz MW, Wintle BA, Broennimann O, Austin M, Ferrier S, Kearney MR, Possingham HP, Buckley YM. 2013. Predicting species distributions for conservation decisions. Ecology Letters 16:1424–1435 DOI 10.1111/ele.12189. Ihlow F, Courant J, Secondi J, Herrel A, Rebelo R, Measey GJ, Lillo F, De Villiers A, Vogt S, De Busschere C, Backeljau T, Rödder D. 2016. Impacts of climate change on the global invasion potential of the African clawed frog Xenopus laevis. PLOS ONE 11(6):e0154869 DOI 10.1371/journal.pone.0154869. Kumschick S, Vimercati G, de Villiers FA, Mokhatla M, Davies SJ, Thorp CJ, Rebelo A, Measey GJ. 2017. Impact assessment with different scoring tools: how well do alien amphibian assessments match? Neobiota 33:53–66 DOI 10.3897/neobiota.33.10376. Lechowicz MJ. 1982. The sampling characteristics of electivity indices. Oecologia 52:22–30 DOI 10.1007/BF00349007.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 14/16 Lobos G, Jaksic FM. 2005. The ongoing invasion of African clawed frogs (Xenopus laevis) in Chile: causes of concern. Biodiversity and Conservation 14:429–439 DOI 10.1007/s10531-004-6403-0. Lobos G, Measey GJ. 2002. Invasive populations of Xenopus laevis (Daudin) in Chile. Herpetological Journal 12:163–168. Lodge DM. 1993. Biological invasions: lessons for ecology. Trends in Ecology & Evolution 8:133–136 DOI 10.1016/0169-5347(93)90025-K. Luiselli LM, Capizzi D, Filippi E, Anibaldi C, Rugiero L, Capula M. 2007. Comparative diets of three populations of an aquatic snake (Natrix tessellata, Colubridae) from mediterranean streams with different hydric regimes. Copeia 2007:426–435 DOI 10.1643/0045-8511(2007)7[426:CDOTPO]2.0.CO;2. McCoid MJ, Fritts TH. 1980. Notes on the diet of a feral population of Xeno- pus laevis (Pipidae) in California. The SouthWestern Naturalist 25:272–275 DOI 10.2307/3671256. Measey GJ. 1998a. Diet of feral Xenopus laevis (Daudin) in South Wales, UK. Journal of Zoology 246:287–298 DOI 10.1111/j.1469-7998.1998.tb00159.x. Measey GJ. 1998b. Terrestrial prey capture in Xenopus laevis. Copeia 3:787–791. Measey GJ, Rödder D, Green SL, Kobayashi R, Lillo F, Lobos G, Rebelo R, Thirion J-M. 2012. Ongoing invasions of the African clawed frog, Xenopus laevis: a global review. Biological Invasions 14:2255–2270 DOI 10.1007/s10530-012-0227-8. Measey GJ, Vimercati G, de Villiers FA, Mokhatla M, Davies SJ, Thorp CJ, Rebelo AD, Kumschick S. 2016. A global assessment of alien amphibian impacts in a formal framework. Diversity and Distributions 22(9):970–981 DOI 10.1111/ddi.12462. Mukherjee A, Williams DA, Wheeler GS, Cuda JP, Pal S, Overholt WA. 2012. Brazilian peppertree (Schinus terebinthifolius) in Florida and South America: evidence of a possible niche shift driven by hybridization. Biological Invasions 14:1415–1430 DOI 10.1007/s10530-011-0168-7. Rehage JS, Barnett BK, Sih A. 2005. Foraging behaviour and invasiveness: do invasive Gambusia exhibit higher feeding rates and broader diets than their noninvasive relatives? Ecology of Freshwater Fish 14:352–360 DOI 10.1111/j.1600-0633.2005.00109.x. Ricciardi A, Rasmussen JB. 1998. Predicting the identity and impact of future biological invaders: a priority for aquatic resource management. Canadian Journal of Fisheries and Aquatic Sciences 55:1759–1765 DOI 10.1890/02-0571. Rödder D, Ihlow F, Courant J, Secondi J, Herrel A, Rebelo R, Measey GJ, Lillo F, De Villiers FA, De Busschere C, Backeljau T, Measey GJ, Lillo F, De Villiers FA, De Busschere C, Backeljau T. Global realized niche divergence in the African-clawed Frog Xenopus laevis. Ecology and Evolution In Press. Rutz C, Whittingham MJ, Newton I. 2006. Age-dependent diet choice in an avian top predator. Proceedings of the Royal Society B: Biological Sciences 273:579–586 DOI 10.1098/rspb.2005.3353. Schafer LN, Platell ME, Valesini FJ, Potter IC. 2002. Comparisons between the influence of habitat type, season and body size on the dietary compositions of fish species in nearshore marine waters. Journal of Experimental Marine Biology and Ecology 278:67–92 DOI 10.1016/S0022-0981(02)00337-4.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 15/16 Schoonbee HJ, Prinsloo JF, Nxiweni JG. 1992. Observations on the feeding habits of larvae, juvenile and adult stages of the African clawed frog, Xenopus laevis, in impoundments in Transkei. Water SA 18:227–236. Schramm M. 1987. Control of Xenopus laevis (Amphibia: Pipidae) in fish ponds with observations on its threat to fish fry and fingerlings. Water SA 13:53–56. Shannon CE, Weaver W. 1964. The mathematical theory of communication. Urbana: University of Illinois Press. Slatyer RA, Hirst M, Sexton JP. 2013. Niche breadth predicts geographical range size: a general ecological pattern. Ecology Letters 16:1104–1114 DOI 10.1111/ele.12140. Soberón J. 2007. Grinnellian and Eltonian niches and geographic distributions of species. Ecology Letters 10:1115–1123 DOI 10.1111/j.1461-0248.2007.01107.x. Stiels D, Gaißer B, Schidelko K, Engler JO, Rödder D. 2015. Niche shift in four non- native estrildid finches and implications for species distribution models. Ibis 157:75–90 DOI 10.1111/ibi.12194. Tillberg CV, Holway DA, Lebrun EG, Suarez AV. 2007. Trophic ecology of invasive Argentine ants in their native and introduced ranges. Proceedings of the National Academy of Sciences of the United States of America 104:20856–20861 DOI 10.1073/pnas.0706903105. Tinsley RC, Stott LC, Viney ME, Mable BK, Tinsley MC. 2015. Extinction of an introduced warm-climate alien species, Xenopus laevis, by extreme weather events. Biological Invasions 17:3183–3195 DOI 10.1007/s10530-015-0944-x. Van Ngo B, Lee YF, Ngo CD. 2014. Variation in dietary composition of granular spiny frogs (Quasipaa verrucospinosa) in central Vietnam. Herpetological Journal 24:245–253. Vanderploeg HA, Scavia D. 1979. Calculation and use of selectivity coefficients of feeding: zooplankton grazing. Ecological Modelling 7:135–149 DOI 10.1016/0304-3800(79)90004-8. Vazquez DP. 2006. Exploring the relationship between niche breadth and invasion success. In: Cadotte MW, McMahon SM, Fukami T, eds. Conceptual ecology and invasions biology. Dordrecht: Springer International Publishing, 317–332. Vogt S, De Villiers FA, Ihlow F, Rödder D, Measey GJ. Competition and feeding ecology in two sympatric Xenopus species (Anura: Pipidae). PeerJ In Press. Wiens JJ, Graham CH. 2005. Niche conservatism: integrating evolution, ecology, and conservation biology. Annual Review of Ecology, Evolution, and Systematics 36:519–539 DOI 10.1146/annurev.ecolsys.36.102803.095431. Wu Z-J, Li Y-M, Wang Y-P. 2007. A comparison of stomach flush and stomach dissection in diet analysis of four frog species. Acta Zoologica Sinica 53:364–372.

Courant et al. (2017), PeerJ, DOI 10.7717/peerj.3250 16/16

APPENDIX D

Characteristics of the ponds used during the field works.

Illustration of the general characteristics of the ponds used during the field works of the PhD. Ponds were generally located in pasture lands, crop cultures or unsued lands.

APPENDIX E

Native amphibians potentially observed during field works.

Native amphibians potentially observed during the field works. From the top left to the bottom right: Triturus cristatus, Lissotriton helveticus, Triturus marmoratus, Bufo spinosus, Rana dalmatina, Pelodytes punctatus, Pelophylax sp. (P. ridibunda or P. lessonae), Hyla arborea, Salamandra salamandra, Epilodea calamita, Alytes obtetricans. During the data collection for chapter II, T. marmoratus, S. salamandra, E. calamita and A. obtetricans have not been detected.

APPENDIX F

Illustration of the changes occurring in a pond during a single year.

Illustration of the changes occurring in a pond during a single year. The picture on the top was taken in april 2014 and the one on the bottom during July of the same year. Between each picture, cattle has occupied the field and used the pond as a trough and a source of nutrients.

APPENDIX G

Courant J, Secondi J, Bereiziat V, Herrel A (2017) Resources allocated to reproduction decrease at the range edge of an expanding population of an invasive anuran. Biol J Linn Soc.

Biological Journal of the Linnean Society, 2017, 122, 157–165. With 3 figures.

Resources allocated to reproduction decrease at the range edge of an expanding population of an invasive amphibian

JULIEN COURANT1*, JEAN SECONDI2,3, VIVIANE BEREIZIAT1 and ANTHONY HERREL1,4

1UMR 7179 CNRS/MNHN, Département d’Ecologie et de Gestion de la Biodiversité, 57 rue Cuvier, Case Postale 55, 75231, Paris Cedex 5, France 2UMR 5023 Écologie des Hydrosystèmes Naturels et Anthropisés, Université Lyon 1, ENTPE, CNRS, Université de Lyon, F-69622 Villeurbanne, France 3UMR 6554 LETG – LEESA, Université d’Angers, 2 Boulevard Lavoisier, 49000 Angers, France 4Ghent University, Evolutionary Morphology of Vertebrates, K. L. Ledeganckstraat 35, B-9000 Gent, Belgium

Received 3 February 2017; revised 11 April 2017; accepted for publication 11 April 2017

Predicting the magnitude and nature of changes in a species’ range is becoming ever more important as an increasing number of species are faced with habitat changes or are introduced to areas outside of the species’ native range. An organism’s investment in life-history traits is expected to change during range shifts or range expansion because populations encounter new ecological conditions. While simulation studies predict that dispersal and reproductive allocation should increase at the range edge, we suggest that reproductive allocation might decrease at the range edge due to energy allocation trade-offs. We studied the reproductive investment of an invasive amphibian, Xenopus laevis, and measured reproductive allocation in three clusters of populations distributed from the centre to the edge of the colonized range of X. laevis in France. Resource allocation was estimated with the scaled mass index of gonads of both sexes during the local period of reproduction of the species. The level of resources allocated to reproduction was lower at the periphery of the colonized range compared to the centre and may be the result of changes in trade- offs between life-history traits. Such a pattern could be explained by interspecific competition or enhanced investment in dispersal capacity.

ADDITIONAL KEYWORDS: invasive species – range expansion – resource allocation – Xenopus laevis.

INTRODUCTION transport and trade is believed to be responsible for the fast rise in the number of invasive species at the Shifts in the distribution of a species occur when range global scale (Lockwood, Hoopes & Marchetti, 2013; edges track environmental changes, when propagules Herrel & van der Meijden, 2014). Analysing the pro- settle beyond an ecological barrier after a discrete cess of range expansion is critical to understand and introduction event or when evolutionary processes predict the dynamics of a species. This knowledge is modify the species’ niche and render novel environ- also important to guide the management of practition- mental conditions suitable. This issue has been paid ers who seek to enhance the expansion of reintroduced renewed attention because of the increasing aware- populations or limit the spread of unwanted aliens (e.g. ness about the scale of global changes (Parmesan & Roman & Darling, 2007; Shine, 2012; Brown, Kelehear Yohe, 2003), including human activities that translo- & Shine, 2013). cate individuals to new areas (Channell & Lomolino, Range expansion typically occurs after the intro- 2000). In this regard, the massive increase in global duction and settlement of a propagule in a novel environment. It is largely driven by population growth, dispersal and density dependence (Fisher, *Corresponding author. E-mail: [email protected] 1937; Skellam, 1951; Burton, Phillips & Travis, 2010).

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These parameters summarize the trade-offs in the MATERIAL AND METHODS life-history traits that influence the probability for an Study area and sampling individual to reproduce, disperse and survive (Burton et al., 2010). Simulation and empirical studies have The African clawed frog was introduced in Western demonstrated that allocation of resources to dispersal France during the late 1980s, and its estimated at the range edge is enhanced during range expan- colonized area was 207 km2 in 2012 (Measey et al., sions (e.g. Thomas et al., 2001; Travis & Dytham, 2012). Contrary to the Mediterranean-like cli- 2002; Hughes, Hill & Dytham, 2003; Simmons & mate of its native range, the population introduced Thomas, 2004; Phillips et al., 2008; Léotard et al., in France occupies an area where the climate is 2009; Phillips, Brown & Shine, 2010; Barton et al., defined as oceanic altered (Joly et al., 2010), with a 2012; Henry, Bocedi & Travis, 2013; Hudson, Brown relatively high annual mean temperature (12.5 °C), & Shine, 2016). Changing dispersal allocation affects and cumulated annual precipitations reaching the trade-offs with other life-history traits that also 800–900 mm, mainly during winter (Joly et al., require energy or resources (Ferriere & Le Galliard, 2010). We sampled ponds with an area of less than 2001). Furthermore, Burton et al. (2010) predicted 500 m2 and a maximum depth between 1 and 3 m. that dispersal and reproductive allocation should Frogs were captured with fykes (60-cm length × increase, whereas resource allocation to competi- 30-cm width × 6-mm mesh diameter). Capture tive ability should be reduced at the range edge. The effort was fixed at 1 trap/50 m2 of water, with a dis- rationale is that long dispersers have a higher prob- tribution of traps as regular as possible inside the ability to reach a new suitable location at the edge pond, during three consecutive nights, and without where density-dependent responses and intra-specific repeating this process more than once in the same competition are low (Brown et al., 2013). At the range pond. Every morning, individuals captured in fykes core, dispersers would mostly reach already colonized were euthanized with a lethal injection of sodium areas (Baguette & Van Dyck, 2007) so that in core pentobarbital and dissected just after death. The areas selection should favour individuals with a high research permit and the authorization for killing competitive ability (Burton et al., 2010). In contrast invasive frogs were provided by the Préfet of the with these predictions, the association of high disper- Deux-Sèvres department. sal capacity and low reproduction has been observed (Hughes et al., 2003). It is therefore important to test whether resource allocation to reproduction is reduced or increased at the range edge of expanding Determination of the optimal period to study populations. reproduction We studied the allocation of resources to reproduc- In California, the period of reproduction of X. laevis tion in an expanding population of the African clawed occurs from January to November, with a 2- to frog, Xenopus laevis (Daudin, 1802) (Anura: Pipidae) 3-month peak during spring (McCoid & Fritts, 1989). in Western France, where it has been introduced. For the French population, the optimal time frame to This species has been used worldwide in laboratories estimate reproductive allocation was determined by for pregnancy tests and as a model in developmen- sampling the gonad mass and the proportion of gravid tal biology and anatomy (Weldon, De Villiers & Du females from March to October 2014 (no individuals Preez, 2007; van Sittert & Measey, 2016). It has been were caught during trapping sessions from November intentionally or unintentionally introduced in several to February). We captured and analysed a total of 408 countries across the world where it is considered as female frogs to determine the breeding period of this invasive (Measey et al., 2012). A study based on spe- population, with monthly captures in six ponds from cies distribution models suggests that the expansion of the range edge, core and intermediate areas (two in the French population is likely to be important during each area), except in October when only two success- the next decades (Ihlow et al., 2016). In another recent ful capture occurred at the range core and the inter- study, an increased allocation to dispersal has been mediate area. New ponds were used every month. The suggested through the observation of an increase in optimal time frame was then defined as the period hind limb morphology and stamina at the range edge when the proportion of gravid females and the rela- of the same population (Louppe, Courant & Herrel, tive mass of reproductive organs [estimated with the 2017). Thus, we hypothesize that at the range edge scaled mass index (SMI), see below] was increasing individuals will allocate fewer resources to reproduc- and when the females are most likely to start breeding tion than at the range core. We test this hypothesis by and have to laid many eggs. We then could assess the estimating the relative mass of the reproductive organ spatial variation of allocation to reproduction on the of males and females with a body condition index. next spring, in 2015.

Estimation of the resource allocation to respectively. For males, 186, 103 and 138 individuals reproduction were captured at the range core, the intermediate area The present study was carried out on three clusters of and the range edge, respectively. Data sampling for the five ponds each, distributed across a part of the colo- reproductive allocation study occurred from 4 April nized range (Fig. 1). The clusters were located at the 2015 to 13 May 2015. core area close to the introduction site of the population, the range edge (18 km from the introduction site) and in an intermediate zone between both areas (11 Morphological and anatomical measures km from the introduction site), respectively. The time Individuals were measured and sexed before dissec- since colonization at the centre of the range is close to tion during 2014 and 2015 data samplings. Males were 30 years (Fouquet, 2001). At the range edge, the study recognized by the presence of reproductive callosities sites were colonized less than 3 years ago, according to on their forelimbs and by the flat shape of their clo- local landscape managers and unpublished data. This aca. Females were identified by the dilated shape of area was selected because it was the most accurately the cloacae and the absence of reproductive callosities identified colonization axis during our field studies on their forelimbs. The combination of these criteria and because no barrier for dispersion was identified avoided misidentification of sex for adults. However, in the landscape. In 2015, 731 frogs were captured to for small individuals [snout–vent length (SVL) study the resource allocation to reproduction, with < 50 mm for males and SVL < 70 mm for females], a three ponds per week frequency (one pond in each sex could not be determined unambiguously based cluster). We captured 82, 102 and 120 females at the on external characters. Individuals measuring less range core, the intermediate area and the range edge, than these threshold values were thus not considered

Figure 1. Location of the three pond clusters of Xenopus laevis sampled to analyse the variation in resource allocation to reproduction. Occurrence data for X. laevis were provided by the Société Herpétologique de France.

as sexually mature and not included in our analysis. the SMI for reproductive organ, gonads, adipose tissues SVL was measured with a digital caliper (Mitutoyo and body mass will be referred in the following sections

Absolute IP67 – precision 0.01 mm). Gonads, the adi- as SMIorgan, SMIgonad, SMIadipose and SMIbody, respectively. pose tissues associated to them and body mass were To determine the best period to study reproductive weighed using a digital balance (XL 2PP – precision allocation and to assess spatial variations, we used the 0.01 g). In order to avoid any measurement bias due to data collected in 2014. We tested the variation of the

the mass of ingested prey items, the stomach content SMIorgan between months using an analysis of variance

was removed from each individual before measuring (ANOVA). Pairwise differences in SMIorgan between body mass. Immature oocytes were detected by the consecutive months were tested using t-tests with an aspect of ovaries. Ovarian stages are numbered from adjustment of α by applying a Bonferroni correction. I to VI. Oocytes at stages I–III are completely plain, With the data collected in 2015, we calculated the whereas oocytes at stages IV–VI are bipolar. Therefore, proportion of gravid females in each cluster (number we considered females as being reproductively active of gravid females/total number of females). Using the from stage IV onwards (Rasar & Hammes, 2006), and same data set, we investigated the spatial variation in we did not use females with oocytes in earlier stages. resource allocation between clusters by comparing the

Reproductive effort is one of the main components of SMIorgan in clusters and by using a generalized linear the reproductive strategy of amphibians (Duellman & mixed model for each sex with cluster type and month Trueb, 1986), and it allows a measurement of the costs as fixed effects and capture site nested within cluster of reproduction for many organisms (Stearns, 1992), as a random effect. We took into account the effect

especially for those that do not provide parental care. of body condition by adding the SMIbody in the linear The role of adipose tissue associated to gonads consists mixed models as a fixed effect. We also provided the

in providing resources to future reproduction events mean values for SMIgonad, SMIadipose and SMIorgan for each during the reproductive season (Girish & Saidapur, cluster in a table. All analyses were carried out using 2000) and is thus an important component of repro- R version 3.0.1 (R Core Team, 2013) and the packages ductive investment. ggplot2 for graphics and nlme (Pinheiro et al., 2013) for the linear mixed models. Statistical analyses To estimate allocation to reproduction, we adjusted the mass of reproductive organs (gonads + adipose tissue) RESULTS by using a body condition index. Most indices found in Determination of the breeding season the literature are based on an ordinary least square In 2014, the SMI was not constant during the (OLS) regression between body size (SVL) and body organ mass. However, those indices do not deal with overd- period of activity of the species (ANOVA, relative ovary mass: F = 2.61, P < 0.05). It was significantly higher ispersion and heteroscedasticity (Peig & Green, 2010). 7,335 Furthermore, it has been recommended not to use OLS in March compared to April (Welch two-sample t-test: t = −6.8993, P < 0.001), but the capture success based indices for studies comparing different popula- 21.28 tions or sites (Jakob, Marshall & Uetz, 1996). We used and the resulting sample size in March were low. The SMI decreased down to the baseline level from July the SMI, which does not present these limitations organ (Peig & Green, 2009). The SMI has already been tested to October (Fig. 2), with a significant increase from April to May (Welch two-sample t-test: t = 2.66, on an anuran, Lithobates catesbeianus, and provided a 172.45 good fit to the data (MacCracken & Stebbings, 2012). P < 0.01), a significant decrease from June to July ^ (Welch two-sample t-test: t = 2.91, P < 0.01) and This index, noted Mi, is calculated according to the fol- 52.44 lowing formula: from September to October (Welch two-sample t-test: t = −11.86, P < 0.0001). There was no significant bSMA 7.12 ^  Lθ  change between March and April (Welch two-sample Mii= M ⋅    Li  t-test: t16.14 = 0.06, P = 0.95), May and June (Welch two- sample t-test: t = −1.43, P = 0.16), July and August where M is the mass of the individual i or the tested 39.45 i (Welch two-sample t-test: t = 0.72, P = 0.48) and organ, L the SVL of the individual, L the mean SVL 111.15 i θ August and September (Welch two-sample t-test: of the sample (for details, see Peig & Green, 2009). The t18.60 = 1.44; P = 0.17). exponent bSMA is the slope of the standard major axis (SMA) regression of M on L. SMI values were estimated based on a and b calculated separately Lθ SMA for each cluster and each sex. This index adjusts the Spatial variation in resource allocation body mass of individuals to the mass they would have In 2015, the proportion of gravid females was 66.26% at the length Lθ . The corrected mass, calculated with for the core, 78.43% for the intermediate and 51.67%

Figure 2. Mean values (± SE) of the SMIorgan of females Xenopus laevis captured in 2014. The number of individuals (N) and

the proportion of gravid females (FGravid) captured each month is mentioned close to the dot of each month.

for the edge cluster. According to the linear mixed mod- of reproduction from the core of the colonized range to els, allocation to reproduction for females decreased its periphery.

from the range core to the intermediate [SMIorgan

(Core) − SMIorgan (Inter) = 1.46] and edge [SMIorgan Determination of the breeding season (Core) – SMIorgan (Edge) = 0.92] ranges (Fig. 3 and

Table 1). In males, SMIorgan was not significantly higher According to our results based on the data collected in (Table 1) in the core cluster than in the intermediate 2014, the proportion of sexually active females varied

cluster [SMIorgan (Core) − SMIorgan (Inter) = 0.34] and between March and October in the French population. significantly higher (Table 1) at the core than at the Reproductive activity peaked for 3 months between

range edge [SMIorgan (Core) – SMIorgan (Edge) = 0.37]. April and June. In South Africa, the breeding period

The decrease in SMIorgan mainly concerned the gonad lasts 3–5 months (Deuchar, 1975). The maximal breed- relative mass for females, whereas, for males, both tes- ing activity lasts only 2–3 months in California, but ticles and adipose tissue masses were reduced at the reproduction can occur from January to November range edge (Table 2). (McCoid & Fritts, 1989). The duration of the optimal breeding period in France and South Africa cannot be strictly compared because Deuchar (1975) only mentioned the breeding period and did not define the DISCUSSION optimal breeding period in itself. In California, water In this study, we observed a modification of the repro- temperature in ponds reaches 20 °C each month of the ductive investment of X. laevis in Western France. This year. Such a high-temperature regime could favour modification consists in a decrease of the relative mass year-round reproduction (McCoid & Fritts, 1989). of reproductive organs during the first part of the peak In France, a drop in temperature could explain the

Figure 3. Reproductive investment of Xenopus laevis in the core, intermediate and edge clusters as represented by the

SMIorgan (mean ± SE) of females (left) and males (right).

Table 1. Results of the linear mixed models performed to test the spatial variation in reproductive allocation across the colonized range of Xenopus laevis in France

Model formula: SMIorg ~ cluster + month + SMIbody+ ~1 |cluster/pond

Effects Estimate SE df t-value P-value

Females Intercept 0.853 0.849 245 1.004 0.316 Cluster core vs. edge −2.359 0.923 23 −2.555 0.018 Cluster core vs. inter −2.427 0.989 23 −2.454 0.022 Month −0.286 0.769 245 −0.272 0.710

SMIbody 0.106 0.009 24 11.989 < 0.001 Random Cluster/pond 1.235 2.844 – – – Males Intercept 0.142 0.010 377 1.364 0.173 Cluster core vs. edge −0.025 0.010 24 −2.384 0.025 Cluster core vs. inter −0.008 0.012 24 −0.701 0.490 Month −0.003 0.006 377 −0.447 0.655

SMIbody 0.003 < 0.001 377 14.380 < 0.001 Random Cluster/pond 0.017 0.035 – – –

We tested a model for each sex, to estimate the variation of SMIorgan according to the cluster, with the month of capture and the SMIbody as covariates. The pond nested in the cluster was used as a random effect.

sharp decrease in the proportion of gravid females and excluded June because most females captured at this the relative gonad mass at the beginning of autumn moment are likely to have already laid some eggs dur- when air temperature usually remains below 20 °C. ing the previous months. Temperature, among others, affects the variation of food abundance, which could be a crucial factor for the sexual activity of the species (Holland & Dumont, 1975; Spatial variation in resource allocation Girish & Saidapur, 2000). In 2015, to study reproduc- In 2015, we found that at the beginning of the opti- tive allocation, we only considered April and May. We mal breeding period, the proportion of gravid females

Table 2. Mean values and SE obtained for each sex for of native competitors (Burton et al., 2010). At low den- the SMI of gonads, adipose tissue and entire reproductive sity of competitors at the range margin, resource alloca- organ in the three clusters tion to competitive ability is minimal, while resources allocated to dispersal and reproduction are maximal. Core Intermediate Edge At higher density of competitors at the range core, allocation to reproduction decreases when competitors Females exhibit high reproductive rate and competitive ability. SMI 5.36 ± 0.56 3.80 ± 0.29 4.09 ± 0.46 gonad This trait has been studied most commonly in inva- SMI 1.02 ± 0.12 0.88 ± 0.10 1.71 ± 0.15 adipose sive plants (Strauss et al., 2002) and is defined as the SMI 6.14 ± 0.56 4.67 ± 0.25 5.22 ± 0.49 organ effect of an individual on resources and its response to Males resource deficiency (Goldberg, 1990). It influences the SMI 0.11 ± 0.004 0.11 ± 0.004 0.08 ± 0.003 gonad survival probability of individuals as a lower invest- SMI 1.22 ± 0.06 0.88 ± 0.05 0.89 ± 0.05 adipose ment in competitive ability leads to a reduced survival SMI 1.33 ± 0.06 0.99 ± 0.05 0.96 ± 0.06 organ in high-density conditions (Burton et al., 2010). In the colonized area, the aquatic habitats used by was lower at the range edge, but the effect was rather X. laevis are occupied by several native amphibian weak. A lower frequency of reproductive females has species that breed during the same period of the year already been observed in an invasive amphibian dur- (Lescure & de Massary, 2012). During field work, we ing range expansion (Hudson et al., 2015). However, observed native amphibians such as Pelophylax sp. in these authors were not able to unambiguously link high densities in every pond. Other species (e.g. Rana this pattern to a lower investment in reproduction at dalmatina, Lissotriton helveticus, Triturus cristatus) the range edge because their data were not initially were often observed at lower densities. Unfortunately, collected for that purpose. Also, in our study, we cannot even if impacts of X. laevis on other amphibians have provide an unambiguous conclusion as the non-gravid already been shown (Lillo, Faraone & Lo Valvo, 2011), females that we caught may have bred earlier in no study has assessed the competition between inva- spring or may reproduce later in the summer. Females sive and native amphibians. Competition through from the core area may also be able to restart ovary trophic resource access is likely to occur (JC et al., pers. development earlier than females from the range edge observ.). Further investigations testing the impact on but laboratory studies remain necessary to test this native amphibians may provide some clarification hypothesis. about this issue. The masses obtained with the SMI appeared much Because females lay eggs in groups of one, two or more informative, with significantly lower values at three eggs (Crump, 2015), it is not possible to tell the range edge compared to the core area for both whether a female has already laid some eggs. We sexes. These modifications concerned the testicles and limited this effect by analysing data collected during adipose tissues for males, both reduced at the range the first 2 months of the peak of reproductive activ- edge. In anurans, this tissue is crucial to allow sper- ity. It is also important to notice that every female is matogenesis in males and the priming of sexual behav- subject to this effect irrespective of the locality where iour (Iela et al., 1979; Grafe, Schmuck & Linsenmair, they are captured. Consequently, this effect should not 1992). At the opposite, no changes in adipose tissue drastically alter the conclusions of this study, but it mass have been detected for females. Its extent is min- should be taken into account in future studies using imal during the breeding season (Iela et al., 1979), and the results from the present study (i.e. care should be there is a negative correlation between adipose tissue taken in interpreting these data as absolute values and ovary mass (Girish & Saidapur, 2000). The lack of of reproductive investment). The pattern of variation differences in adipose tissue mass for females between in reproductive organ mass that we observed is pri- the range edge and the core area could be explained by marily expected to reflect changes in the allocation of the time of the study, that is the first 2 months of the resources to different life-history traits. However, the peak of reproductive activity. For both sexes, stronger phenomenon called gene surfing, that is the fact that differences are expected when the development of the mutations occurring on the range edge travel with the adipose tissue is maximal. wave front and induce mutant populations (Edmonds, The results of our study are corroborated by the Lillie & Cavalli-Sforza, 2004), could provide an alter- increased allocation to dispersal observed in the same native explanation. This process may sort a different invasive population of X. laevis (Louppe et al., 2017). pool of alleles in different directions as propagules A study using individual-based models suggested that are loosely connected at the margins of an expanding resource allocation may change depending on the pres- range (Excoffier et al., 2009). Genetic analyses may ence, the reproductive rate and the competitive ability help to solve this issue. Sampling transects pointing

away from the range core in different directions would Channell R, Lomolino MV. 2000. Dynamic biogeography and certainly also help to provide more robust conclusions. conservation of endangered species. Nature 403: 84–86. Finally, common garden studies on resource allocation Crump M. 2015. Anuran reproductive modes: evolving per- in expanding populations have already demonstrated spectives. Journal of Herpetology 49: 1–16. that heritability could have an effect on the observed Deuchar EM. 1975. Xenopus: the South African clawed frog. pattern for a life-history trait (Brown, Phillips & Shine, London: John Wiley and Sons. 2014). The African clawed frog is an interesting model Duellman W, Trueb L. 1986. Biology of amphibians. to address this issue because the ecological condi- Baltimore: Johns Hopkins University Press. tions (climate, communities) encountered in its native Edmonds C, Lillie AS, Cavalli-Sforza LL. 2004. Mutations arising in the wave front of an expanding population. range and the colonized areas differ. This species, by Proceedings of the National Academy of Sciences of the offering intra-specific comparisons across populations United States of America 101: 975–979. introduced independently in different countries, may Excoffier L, Foll M, Petit RJ. 2009. Genetic consequences of bring insights into the processes of range expansion range expansions. Annual Review of Ecology, Evolution, and and invasion. Systematics 40: 481–501. Ferriere R, Le Galliard JF. 2001. Invasion fitness and the adaptive dynamics of space related traits. In: Clobert J, ACKNOWLEDGEMENTS Danchin E, Dhondt AA, Nichols JD, eds. Dispersal. Oxford: Oxford University Press, 57–79. We gratefully thank the two anonymous reviewers of Fisher RA. 1937. The wave of advance of advantageous genes. this study and Prof. John A. Allen who led the reviews Annals of Human Genetics 7: 355–369. of our study. We thank the Agglomération du Bocage Fouquet A. 2001. Des clandestins aquatiques. Zamenis 6: Bressuirais and the Communauté de Communes du 10–11. Thouarsais for their help during the fieldworks. We Girish S, Saidapur SK. 2000. Interrelationship between food also thank the Société Herpétologique de France for availability, fat body, and ovarian cycles in the frog, Rana providing us the occurrence data of Xenopus laevis. tigrina, with a discussion on the role of fat body in anuran This research was funded by the ERA-Net BiodivERsA, reproduction. The Journal of Experimental Zoology 286: with the national funders ANR, DFG, BELSPO and 487–493. FCT, as part of the 2013 BiodivERsA call for research Goldberg DE. 1990. Components of resource competition in proposals. INVAXEN ‘Invasive biology of Xenopus lae- plant communities. In: Grace J, Tilman D, eds. Perspectives vis in Europe: ecology, impact and predictive models’ on plant competition. San Diego: Academic Press, 27–49. Grafe TU, Schmuck R, Linsenmair KE. 1992. Reproductive project ANR-13-EBID-0008-01. AH, JS and JC con- energetics of the African reed frogs, Hyperolius viridifla- ceived the study. JC and VB collected data in the field. vus and Hyperolius marmoratus. Physiological Zoology 65: JC, VB and JS analysed data, and all authors contrib- 153–171. uted to the writing of the paper. Henry RC, Bocedi G, Travis JM. 2013. Eco-evolutionary dynamics of range shifts: elastic margins and critical thresholds. Journal of Theoretical Biology 321: 1–7. REFERENCES Herrel A, van der Meijden A. 2014. An analysis of the live reptile and amphibian trade in the USA compared to the Baguette M, Van Dyck H. 2007. Landscape connectivity and global trade in endangered species. Herpetological Journal animal behavior: functional grain as a key determinant for 24: 103–110. dispersal. Landscape Ecology 22: 1117–1129. Holland CA, Dumont JN. 1975. Oogenesis in Xenopus laevis Barton KA, Hovestadt T, Phillips BL, Travis JM. 2012. (Daudin). Cell and Tissue Research 162: 177–184. Risky movement increases the rate of range expansion. Hudson CM, Brown GP, Shine R. 2016. Athletic anurans: Proceedings of the Royal Society of London B: Biological the impact of morphology, ecology and evolution on climb- Sciences 279: 1194–1202. ing ability in invasive cane toads. Biological Journal of the Brown GP, Kelehear C, Shine R. 2013. The early toad Linnean Society 119: 992–999. gets the worm: cane toads at an invasion front benefit from Hudson CM, Phillips BL, Brown GP, Shine R. 2015. higher prey availability. The Journal of Animal Ecology 82: Virgins in the vanguard: low reproductive frequency in 854–862. invasion-front cane toads. Biological Journal of the Linnean Brown GP, Phillips BL, Shine R. 2014. The straight and Society 116: 743–747. narrow path: the evolution of straight-line dispersal at a Hughes CL, Hill JK, Dytham C. 2003. Evolutionary trade- cane toad invasion front. Proceedings of the Royal Society of offs between reproduction and dispersal in populations London B: Biological Sciences 281: 1–7. at expanding range boundaries. Proceedings of the Royal Burton OJ, Phillips BL, Travis JM. 2010. Trade-offs and the Society of London B: Biological Sciences 270: S147–S150. evolution of life-histories during range expansion. Ecology Iela L, Milone M, Caliendo MF, Rastogi RK, Chieffi G. 1979. Letters 13: 1210–1220. Role of lipids in the physiology of the testis of Rana esculenta:

annual changes in the lipid and protein content of the liver, Peig J, Green AJ. 2010. The paradigm of body condition: a fat body, testis and plasma. Bolletino di Zoologia 46: 11–16. critical reappraisal of current methods based on mass and Ihlow F, Courant J, Secondi J, Herrel A, Rebelo R, Measey length. Functional Ecology 24: 1323–1332. GJ, Lillo F, De Villiers FA, Vogt S, De Busschere C, Phillips BL, Brown GP, Shine R. 2010. Life-history evolu- Backeljau T, Rödder D. 2016. Impacts of climate change tion in range-shifting populations. Ecology 91: 1617–1627. on the global invasion potential of the African clawed frog Phillips BL, Brown GP, Travis JMJ, Shine R. 2008. Reid’s Xenopus laevis. PLoS ONE 11: e0154869. paradox revisited: the evolution of dispersal kernels during Jakob EM, Marshall SD, Uetz GW. 1996. Estimating fitness: range expansion. The American Naturalist 172: S34–S48. a comparison of body condition indices. Oikos 77: 61–67. Pinheiro J, Bates D, DebRoy S, Sarkar D. 2013. nlme: lin- Joly D, Brossard T, Cardot H, Cavailhès J, Hilal M, ear and nonlinear mixed effects models. 1–3. Wavresky P. 2010. Les types de climats en France, une con- R Core Team. 2013. R: a language and environment for sta- struction spatiale. Cybergeo: European Journal of Geography, tistical computing. Vienna: R Foundation for Statistical 501. Avaiable at: http://cybergeo.revues.org/23155 Computing. Léotard G, Debout G, Dalecky A, Guillot S, Gaume L, Rasar MA, Hammes SR. 2006. The physiology of the Xenopus McKey D, Kjellberg F. 2009. Range expansion drives dis- laevis ovary. In: Clifton NJ, ed. Methods in molecular biology. persal evolution in an equatorial three-species symbiosis. Totowa: Humana Press, 17–30. PLoS ONE 4: e5377. Roman J, Darling JA. 2007. Paradox lost: genetic diversity Lescure J, de Massary JC. 2012. Atlas des amphibiens et and the success of aquatic invasions. Trends in Ecology & reptiles de France. Biotope: Muséum National d’Histoire Evolution 22: 454–464. Naturelle. Shine R. 2012. Invasive species as drivers of evolution- Lillo F, Faraone FP, Lo Valvo M. 2011. Can the introduc- ary change: cane toads in tropical Australia. Evolutionary tion of Xenopus laevis affect native amphibian populations? Applications 5: 107–116. Reduction of reproductive occurrence in presence of the inva- Simmons AD, Thomas CD. 2004. Changes in dispersal dur- sive species. Biological Invasions 13: 1533–1541. ing species’ range expansions. The American Naturalist 164: Lockwood JL, Hoopes MF, Marchetti MP. 2013. Invasion 378–395. ecology. Oxford: John Wiley and Sons. van Sittert L, Measey GJ. 2016. Historical perspectives on Louppe V, Courant J, Herrel A. 2017. Differences in mobil- global exports and research of African clawed frog (Xenopus ity at the range edge of an expanding invasive population laevis). Transactions of the Royal Society of South Africa 71: of Xenopus laevis in the west of France. The Journal of 157–166. Experimental Biology 220: 278–283. Skellam JG. 1951. Random dispersal in theoretical popula- MacCracken JG, Stebbings JL. 2012. Test of a body con- tions. Biometrika 38: 196–218. dition index with Amphibians. Journal of Herpetology 46: Stearns SC. 1992. The evolution of life histories. New York: 346–350. Oxford University Press. McCoid MJ, Fritts TH. 1989. Growth and fatbody cycles in Strauss SY, Rudgers J, Lau J, Irwin RE. 2002. Direct and feral populations of the African clawed frog, Xenopus laevis ecological costs of resistance to herbivory. Trends in Ecology (Pipidae), in California with comments on reproduction. The and Evolution 17: 278–285. Southwestern Naturalist 34: 499–505. Thomas CD, Bodsworth EJ, Wilson RJ, Simmons AD, Measey GJ, Rödder D, Green SL, Thirion J-M. 2012. Davies ZG, Musche M, Conradt L. 2001. Ecological and Ongoing invasions of the African clawed frog, Xenopus laevis: evolutionary processes at expanding range margins. Nature a global review. Biological Invasions 14: 2255–2270. 411: 577–581. Parmesan C, Yohe G. 2003. A globally coherent fingerprint of Travis JMJ, Dytham C. 2002. Dispersal evolution during climate change impacts across natural systems. Nature 421: invasions. Evolutionary Ecology Research 4: 1119–1129. 37–42. Weldon C, De Villiers A, Du Preez LH. 2007. Quantification Peig J, Green AJ. 2009. New perspectives for estimating body of the trade in Xenopus laevis from South Africa, with impli- condition from mass/length data: the scaled mass index as an cations for biodiversity conservation. African Journal of alternative method. Oikos 118: 1883–1891. Herpetology 56: 77–83.

APPENDIX H

Louppe V, Courant J, Herrel A (2017) Differences in mobility at the range edge of an expanding invasive population of Xenopus laevis in the West of France? J Exp Biol 220:278–283.

RESEARCH ARTICLE Differences in mobility at the range edge of an expanding invasive population of Xenopus laevis in the west of France Vivien Louppe, Julien Courant and Anthony Herrel*

ABSTRACT subsequently predicted increased dispersal and reproductive rates Theoretical models predict that spatial sorting at the range edge of along the edge of the expansion range (Hallatschek and Nelson, expanding populations should favor individuals with increased 2007; Excoffier and Ray, 2008; Excoffier, 2009; Burton et al., 2010; mobility relative to individuals at the center of the range. Despite the Travis et al., 2010). Lower predation pressure by specialist predators fact that empirical evidence for the evolution of locomotor and reduced competition (Burton et al., 2010; Phillips et al., 2010; performance at the range edge is rare, data on cane toads support Brown et al., 2013), in addition to an increased risk of kin this model. However, whether this can be generalized to other competition resulting from low population density (Kubisch et al., species remains largely unknown. Here, we provide data on 2013), may also encourage an increased dispersal rate at the locomotor stamina and limb morphology in individuals from two colonization front. However, processes such as mutation surfing sites: one from the center and one from the periphery of an expanding (Travis et al., 2010), Allee effects (Travis and Dytham, 2002) or population of the clawed frog Xenopus laevis in France where it was allocation trade-offs (Bishop and Peterson, 2006; Fronhofer and introduced about 30 years ago. Additionally, we provide data on the Altermatt, 2015) may prevent this from happening. Thus, morphology of frogs from two additional sites to test whether the predictions from theoretical models may not always hold and observed differences can be generalized across the range of this these predictions remain to be tested. The increase of reproductive species in France. Given the known sexual size dimorphism in this output at the range edge, for example, remains controversial species, we also test for differences between the sexes in locomotor (Hughes et al., 2003; Karlsson and Johansson, 2008; Bonte et al., performance and morphology. Our results show significant sexual 2012; Hudson et al., 2015). However, the allocation of resources to dimorphism in stamina and morphology, with males having longer dispersal has been relatively well documented. For example, the fast legs and greater stamina than females. Moreover, in accordance with dispersal rate and associated phenotypic traits that are observed in the predictions from theoretical models, individuals from the range vanguard populations of the invasive cane toad Rhinella marina in edge had a greater stamina. This difference in locomotor Australia (Brown et al., 2007; Alford et al., 2009; Phillips et al., performance is likely to be driven by the significantly longer limb 2008) provide a nice illustration of a dispersal phenotype in a segments observed in animals in both sites sampled in different areas rapidly expanding population. along the range edge. Our data have implications for conservation The present study focuses on another highly and globally because spatial sorting on the range edge may lead to an accelerated invasive amphibian, the African clawed frog Xenopus laevis Daudin increase in the spread of this invasive species in France. 1802. The use of X. laevis as a model system in developmental and cellular biology (Gurdon and Hopwood, 2000) has resulted in the KEY WORDS: Frog, Locomotion, Invasion, Stamina presence of this species in laboratories world-wide. Invasive populations of X. laevis have since become established globally as INTRODUCTION a result of accidental as well as voluntary releases from research Dispersal, typically defined as a permanent movement away from a facilities and through the release of animals from the pet trade site of birth and/or reproduction (Clobert et al., 2009), may be due to (Measey et al., 2012). Despite a growing body of literature on the discrete and repeated introduction of propagules beyond ecological invasion range and the impacts of this species on autochthonous barriers. Alternatively, evolutionary mechanisms that make new ecosystems (Lafferty and Page, 1997; Lillo et al., 2005, 2011; Lobos ecological niches accessible may also spur dispersal (Wilson et al., and Jaksic, 2005; Eggert and Fouquet, 2006; Fouquet and Measey, 2009). Dispersal is dependent on many factors including an 2006; Robert et al., 2007; Faraone et al., 2008; Rebelo et al., 2010; organism’s phenotype, sex, age, reproductive output, the intensity Measey et al., 2012; De Busschere et al., 2016), this species has of competition and environmental conditions (Stevens et al., 2010). never been used to test the predictions of dispersal models. Our The synergistic influence of these parameters may result in the study focuses on an invasive population of X. laevis in the west of spatial differentiation of expanding populations (Shine et al., 2011). France. Its introduction has been suggested to be associated with the Spatial sorting and lower population density have been documented presence of a research laboratory where Xenopus were bred and at the invasion front of expanding populations (e.g. Phillips et al., maintained until the facility closed in the early 1980s (Fouquet and 2010). Theoretical models based on these observations have Measey, 2006). Animals were officially first reported in 1998 when they were observed in few ponds around the likely site of introduction. However, residents of the area subsequently UMR 7179 C.N.R.S./M.N.H.N., Département d’Ecologie et de Gestion de la Biodiversité, 57 rue Cuvier, Case postale 55, Paris Cedex 5 75231, France. suggested that animals had been in these ponds since the early 1980s. Since then, animals have been expanding at a steady rate and *Author for correspondence ([email protected]) they now occupy an area of over 2000 square kilometers. The aim of A.H., 0000-0003-0991-4434 the present study was to test whether X. laevis at the range edge show evidence of dispersal phenotypes. Specifically, we test

Received 21 July 2016; Accepted 26 October 2016 whether animals at the range edge have longer limbs and greater Journal of Experimental Biology

278 RESEARCH ARTICLE Journal of Experimental Biology (2017) 220, 278-283 doi:10.1242/jeb.146589 locomotor performance than animals near the likely site of release. periphery and 46 from the center; sites 3 and 4, respectively, in Fig. 1) To investigate this, we analyzed terrestrial endurance capacity and were killed in the field using an overdose of anesthetic (MS222) limb morphology for individuals from the center and the periphery according to institutional guidelines, preserved in formaldehyde and of the range. used for morphometric analyses.

MATERIALS AND METHODS Morphometrics All individuals (N=164; 84 from the periphery and 80 from the center All individuals were weighed (Ohaus, Brooklyn, NY, USA; of the range) were caught in ponds and bodies of standing water precision±0.1 g) and measured using a digital caliper (Mitutyo; within their current range using fykes. The range of X. laevis in precision±0.01 mm). Body dimensions were measured following western France is identified through regular monthly trapping Herrel et al. (2012). A summary of the morphometric data is campaigns by local fish and wildlife officers and currently covers provided in Table 1. three departments (Vienne, Deux-Sevres̀ and Maine-et-Loire) and an area of ∼2000 km2 (Fig. 1). Two pairs of sites were used in this study: Performance one site at the center of the range, near the introduction point, and one Stamina tests were performed at 22°C, which is considered the site at the periphery. For each site, all individuals were caught in a optimal temperature for the species (Casterlin and Reynolds, 1980; single pond. Individuals from the first pair (N=87; 53 from the Miller, 1982). Animals (sites 1 and 2) were placed in individual periphery and 34 from the center; sites 1 and 2, respectively, in Fig. 1) containers with some water for 1 h in an incubator set at 22°C prior were caught, brought back and housed at the Function and Evolution to each test. The body temperature (Tb) of each individual was (FUNEVOL) laboratory at the Muséum National d’Histoire recorded using a K-type thermocouple before and after stamina trial Naturelle in Paris, France. Specimens were housed in groups of 6– as the room temperature was slightly lower than the temperature of 10 individuals in 50 liter aquaria at a temperature of 23°C and fed the incubator (19°C) causing the animals’ body temperature to drop with beef heart and mosquito larvae. All individuals were pit-tagged slightly during the trials. Between trials, animals were returned to (Nonatec, Lutonic International, Rodange, Luxembourg), allowing their aquaria, fed, and allowed to rest for at least 24 h. Three trials unambiguous identification of each individual during study. per individual were performed and only the single best trial was Individuals from the second pair of sites (N=77; 31 from the retained for further analysis. Stamina was measured by chasing each

4 La Loire

Saumur

Chemillé-Melay

Doué-la-Fontaine

Vihiers

Le Layon e

La Div

1 3

Le Thouet

N NW NE Thouars W E

SW SE S L’Argenton 2

Fig. 1. Current distribution of the invasive population of Xenopus laevis in the west of France. Indicated are the point of introduction and the four sites used in this study. Small dots indicate ponds where X. laevis are present. Journal of Experimental Biology

279 RESEARCH ARTICLE Journal of Experimental Biology (2017) 220, 278-283 doi:10.1242/jeb.146589

Table 1. Morphometric traits (means±s.e.) measured for each site and sex Sex Site 1 (center) Site 2 (periphery) Site 3 (center) Site 4 (periphery) SVL (mm) Female 83.37±1.03 65.01±1.03 86.70±1.04 80.54±1.04 Male 70.15±1.04 59.84±1.02 69.50±1.03 63.53±1.04 Femur length (mm) Female 27.93±1.02 28.84±1.01 27.73±1.02 27.80±1.01 Male 29.65±1.02 30.34±1.01 30.41±1.02 31.48±1.02 Tibia length (mm) Female 24.32±1.03 26.24±1.02 24.38±1.08 25.23±1.05 Male 27.35±1.03 26.61±1.02 25.59±1.07 23.71±1.07 Astragalus length (mm) Female 15.67±1.03 17.82±1.02 14.55±1.03 15.45±1.02 Male 18.54±1.03 18.28±1.02 15.14±1.03 15.31±1.03 Longest toe length (mm) Female 26.73±1.02 26.67±1.02 30.06±1.26 31.55±1.01 Male 29.11±1.02 28.12±1.02 31.05±1.02 32.66±1.02 Humerus length (mm) Female 10.59±1.04 11.86±1.03 11.22±1.32 12.02±1.19 Male 13.49±1.04 13.21±1.03 20.51±1.29 13.15±1.29 Radius length (mm) Female 9.89±1.04 9.68±1.03 12.05±1.03 13.58±1.02 Male 12.13±1.04 11.80±1.03 11.80±1.03 14.39±1.03 Hand length (mm) Female 4.00±1.04 4.31±1.03 4.52±1.05 5.16±1.03 Male 4.41±1.03 4.38±1.03 4.41±1.05 5.19±1.05 Longest finger length (mm) Female 10.02±1.03 9.35±1.02 11.25±1.03 12.42±1.02 Male 11.04±1.03 10.35±1.02 10.84±1.03 12.94±1.03 Ilium width (mm) Female 15.78±1.02 16.63±1.02 15.07±1.03 15.10±1.02 Male 15.17±1.02 16.07±1.02 15.31±1.02 15.59±1.03 Mass (g) Female 29.51±1.03 36.14±1.03 28.84±1.06 26.915±1.04 Male 30.13±1.03 36.64±1.02 27.54±1.06 25.76±1.06 Site 1, N=14 males and N=20 females; site 2, N=33 males and N=20 females; site 3, N=31 males and N=15 females; site 4, N=20 males and N=20 females. individual down a 3 m long circular track covered with cork. to preservation of the animals (sites 3 and 4). An ANOVA was Animals were chased until exhaustion, which was identified by the performed to test for differences in the body temperature of the lack of a righting response. Note that individuals recovered quickly animals from sites 1 and 2 after the trials. Given that both SVL (sites from these trials and were immediately ready to eat when placed 1 and 2: F1,83=47.92; P<0.01; sites 3 and 4: F1,73=6.35; P=0.01) back in their home aquaria. For each individual, we recorded both and body temperature (F1,83=20,65; P<0.001) were different the total distance covered and time spent moving until exhaustion. between animals from different sites they were incorporated as Statistical analyses were performed using the maximum distance covariates in our multivariate analyses. Next, a multivariate analysis covered and the maximum time spent moving for each individual (MANCOVA), with SVL and body temperature as covariates, was out of the three trials (Table 2). performed to test whether stamina, identified here as the maximum time and the maximum distance moved until exhaustion, differed Statistical analyses between sexes and sites. All analyses were performed using SPSS All data were log10 transformed to meet assumptions of normality v.22 (IBMSPSS, Chicago, IL, USA). and homoscedasticity. To test for differences in size [snout–vent length (SVL)] between sexes, and between center and edge sites, RESULTS univariate analyses (ANOVAs) were performed. Differences in Snout–vent length was significantly different between males and body mass, the morphology of the ilium and limb dimensions were females from sites 1 and 2 (F1,83=18.80; P<0.01), with females tested between sexes and populations using multivariate ANOVAs being on average 16% larger than males. Similarly, females were on with the SVL as a covariate (MANCOVA). These analyses were average 20% larger when comparing preserved animals from sites 3 performed independently within each pair of sites (comparison of and 4 (F1,73=50.32; P<0.01; Table 3). SVL was also significantly site 1 with site 2, and site 3 with site 4) to avoid potential biases due different between populations, with individuals from the center being larger than individual from the periphery (sites 1 and 2: F1,83=47.92; P<0.01; sites 3 and 4: F1,73=6.35; P=0.01). Table 2. Mean performance trait and body temperature at the end of the Individuals from site 1 (center of the range) are on average 20% trial larger than those from site 2 (periphery) (Table 1, Fig. 2). Population Sex Mean±s.e. Individuals from site 3 (center of the range) are on average 8% larger than those from site 4 (periphery; Table 1, Fig. 2). Tb (°C) Center Female 22.28±1.01 Male 21.83±1.01 Periphery Female 20.89±1.01 Male 21.63±1.01 Table 3. Results of univariate analyses testing for differences in SVL Distance (cm) Center Female 1241.65±1.12 and body temperature at the end of the stamina trial between Male 1137.63±1.12 populations and sexes Periphery Female 1300.17±1.11 Variable Source F d.f. Error P-value Male 1721.87±1.09 Time (s) Center Female 75.51±1.12 SVL Population 47.92 1 83 <0.01 Male 91.62±1.11 Sex 18.80 1 83 <0.01 Periphery Female 110.15±1.10 Population×sex 2.18 1 83 0.14 Male 148.59±1.08 Tb Population 20.65 1 83 <0.01 Sex 0.88 1 83 0.35 Center, N=14 males and N=20 females; periphery, N=33 males and N=20 Population×sex 12.51 1 83 <0.01 females. Journal of Experimental Biology

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Females Males Females Males Table 5. MANCOVA performed on the morphometric data from sites 3 and 4 with SVL as covariate 90 Wilks’ Effect Variable lambda F d.f. Error P-value

85 Population 0.59 2.88 14 59 <0.01 Femur 38.68 1 72 <0.01 Sex 0.48 4.66 14 59 <0.01 Toe 4.34 1 72 0.04 80 Radius 23.61 1 72 <0.01 Hand 7.53 1 72 <0.01 Finger 16.03 1 72 <0.01 75 Population×sex 0.78 1.19 14 59 0.31

length are significantly longer in males compared with females SVL (mm) SVL 70 (Table 1). Limb dimensions were significantly different between populations as well (sites 1 and 2: Wilks’ lambda=0.61; ’ 65 F11,72=4.25; P<0.01; Table 4; sites 3 and 4: Wilks lambda=0.48; F14,59=4.62; P<0.01; Table 5). Individuals from site 2 (the range edge) have significantly longer astragali (on average 5% longer), a 60 wider ilium (on average 5% wider) and a higher body mass (on average 18% heavier) than individuals from site 1 (center). Individuals from site 4 (periphery) have significantly longer femurs 55 (on average 10% longer; Tables 1, 5) than individuals from site 3 (center). y) y) Stamina, with body length and temperature as covariates, is significantly different between males and females (Wilks’ lambda=0.91; F2,80=4.11; P<0.02), with males being capable of Site 1 (center) Site 1 (center) Site 3 (center) Site 3 (center) moving an average of 23% longer for a given body size and Site 2 (periphery) Site 2 (peripher Site 4 (periphery) Site 4 (peripher temperature (Table 6). Stamina is also significantly different ’ Fig. 2. Snout–vent length for males and females from sites at the center between individuals from the center and the periphery (Wilks and the periphery of the range. Values are means±s.d. Site 1, N=14 males lambda=0.83; F2,80=8.09; P<0.01), with individuals from the range and N=20 females; site 2, N=33 males and N=20 females; site 3, N=31 males edge moving 35% longer before exhaustion (Fig. 3, Table 6). and N=15 females; site 4, N=20 males and N=20 females. Moreover, the distance moved also showed a trend for animals from the periphery to move a greater distance for a given body size Size-corrected limb dimensions were also significantly different compared with animals from the center of the range (P=0.06; between males and females from sites 1 and 2 (Wilks’ lambda=0.53; Table 6). F11,72=5.84; P<0.01; Table 4), and sites 3 and 4 (Wilks’ lambda=0.59; F14,59=2.88; P<0.01; Table 5). Within sites 1 and DISCUSSION 2, all morphological traits except for hand length were significantly Our results show significant differences in stamina and limb greater for males (Table 4). Male forelimbs are on average 15% morphology for two sites in the range of X. laevis in France, one longer and hind limbs 7% longer than those of females of a given from the center and one on the edge of the range. Individuals at the body size (Table 1). For sites 3 and 4, the toe, radius, hand and finger range edge showed greater stamina and had longer legs. Our analyses of the limb morphology in a second set of populations also show longer limbs for animals from the range edge, suggesting that Table 4. MANCOVA performed on the morphometric data from sites 1 this is a more general phenomenon. However, measurements of and 2 with SVL as covariate locomotor performance in frogs from additional sites are needed to Wilks’ better understand whether the longer limbs observed in animals Effect Variable lambda F d.f. Error P-value from these additional sites also result in differences in endurance Population 0.61 4.25 11 72 <0.01 capacity. Mass 40.48 1 82 <0.01 Astragalus 4.56 1 82 0.04 Ilium width 6.00 1 82 0.02 Table 6. MANCOVA performed on the performance traits with SVL and Finger 6.55 1 82 0.01 temperature as covariates Sex 0.53 5.84 11 72 <0.01 Wilks’ Femur 11.37 1 82 <0.01 Effect Variable lambda F d.f. Error P-value Tibia 7.71 1 82 0.01 Astragalus 16.75 1 82 >0.01 Population 0.83 8.09 2 80 <0.01 Longest toe hind 11.49 1 82 >0.01 Distance 3.74 1 81 0.06 Humerus 26.00 1 82 >0.01 Time 15.08 1 81 <0.01 Radius 33.09 1 82 >0.01 Sex 0.91 4.11 2 80 0.02 Longest toe front 18.56 1 82 >0.01 Distance 0.90 1 81 0.35 Hand 2.43 1 82 0.12 Time 6.75 1 81 0.01 Population×sex 0.79 1.73 11 72 0.08 Population×sex 0.95 2.27 2 80 0.11 Journal of Experimental Biology

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160 (Herrel et al., 2012) and R. marina (Phillips et al., 2006). The length of the foot could also play a key role in aquatic locomotion, because foot size and rotation are crucial for the generation of thrust in aquatic frogs (Richards, 2010). The ilium may also play an 140 important role in aquatic locomotion, as it has been suggested that the sliding of sacral vertebrae along the ilia during swimming improves swimming speed in Xenopus frogs (Videler and Jorna, 1985). Therefore, in addition to having a better terrestrial locomotor 120 endurance, frogs from the range edge may also have an improved swimming performance compared with individuals from the center of the range. However, this remains to be tested for individuals from the four sites included in this study. Time of effort (s) of effort Time 100 From the perspective of conservation biology, it is of particular importance to pay attention to the rapid evolution of morphological and physiological traits observed in X. laevis. As highlighted in the 80 case of the invasion of R. marina in Australia (Brown et al., 2007; Alford et al., 2009; Phillips et al., 2006, 2008), the fast optimization of dispersal abilities can lead to situations that can be unmanageable for conservation biologists. Our study demonstrates an improved Female Male Female Male terrestrial locomotor performance at the range edge but also Site 1 (center) Site 2 (periphery) suggests better locomotor abilities in an aquatic environment. In order to optimize the conservation efforts and the preservation of Fig. 3. Mean time until exhaustion for males and females from autochthonous ecosystems, a better understanding of the populations at the center versus the periphery of the range. Values are means±s.d. Site 1, N=14 males and N=20 females; site 2, N=33 males and physiological, evolutionary and behavioral responses of invasive N=20 females. species that can impact dispersal and colonization is key.

CONCLUSION In addition, our results highlight morphological and locomotor This study showed significant differences in performance and differences between males and females in X. laevis. Females are morphology in X. laevis from sites in the center versus the significantly larger than males (Fig. 2, Table 2) as is common in periphery of the range. Moreover, these differences have evolved frogs and in Xenopus species more specifically (Zug, 1978; Measey since their introduction less than 40 years ago. This suggests and Tinsley, 1998; Herrel et al., 2012). Males also possess relatively that, as in other invasive amphibians, spatial sorting has resulted longer forelimbs and hind limbs than females (Table 2). in the evolution of locomotor capacity, improving the dispersal Furthermore, males show a better endurance capacity (specifically ability of individuals on the range edge. Although experiments the time until exhaustion) than females (Fig. 3, Table 6) similar to are needed to test the genetic basis of these differences, the fact observations for Xenopus tropicalis (Herrel et al., 2012). The greater that there is more than 15 km between the sites from the center endurance capacity in males relative to females may be explained by to the periphery suggests that gene flow may be limited and thus their relatively larger limbs and lower body mass allowing them to these subpopulations may have diverged significantly. Finally, keep moving longer than females for their body size. This could be our results are consistent with models predicting the allocation of beneficial for males during courting and reproduction. The longer resources to dispersal at the range edge of expanding forelimbs observed in males may additionally provide males an populations. However, it remains to be tested whether this advantage during mating, allowing them to maintain their grasp on implies trade-offs with other traits such as reproductive females during amplexus (Measey and Tinsley, 1997). investment, immunity or competitive ability. In addition to the differences between sexes, our data also show significant differences in body size and endurance capacity between Acknowledgements ́ the center and the periphery of the range (Fig. 3, Table 6). The authors would like to thank the communaute des communes Thouarsais for their help in the identification of field sites and three anonymous reviewers and the Individuals from the edge of the expansion range are smaller and editor for constructive and helpful comments on our manuscript. have a higher endurance than those from the center of the range. Our results also highlight differences in body dimensions, particularly in Competing interests limb segments involved in locomotion, such as the astragalus and The authors declare no competing or financial interests. ilium for individuals from site 2, and the femur for individuals from site 4. The difference in the specific skeleton elements impacted is Author contributions A.H. devised the study; J.C. captured animals in the field; V.L. performed the intriguing and may be due to specific differences between locomotor trials; V.L. and A.H. analyzed the data; all authors contributed to the populations. Whereas individuals from site 2 got to this locality writing of the manuscript. by overland migration, animals from site 4 may have been taking advantage of waterways to reach their current site. Although the Funding limb segments involved are different, these results suggest a Funded by a BiodivERsA ERA-NET project entitled ‘Invasive biology of Xenopus ’ common response: hind limbs are longer in individuals from the laevis in Europe: ecology, impact and predictive models (INVAXEN). periphery, which is likely to enhance their locomotor capacity. References Indeed, previous studies have demonstrated a relationship between Alford, R. A., Brown, G. P., Schwarzkopf, L., Phillips, B. and Shine, R. (2009). limb dimensions and locomotor performance. Higher endurance Comparisons through time and space suggest rapid evolution of dispersal abilities are attributed to relatively longer hind limbs in X. tropicalis behaviour in an invasive species. Wildl. Res. 36, 23-28. Journal of Experimental Biology

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Bishop, M. J. and Peterson, C. H. (2006). When r-selection may not predict Lafferty, K. D. and Page, C. J. (1997). Predation on the endangered tidewater goby, introduced-species proliferation: predation of a non-native oyster. Ecol. Appl. 16, Eucyclogobius newberryi, by the introduced African clawed frog, Xenopus laevis, 718-730. with notes on the frog’s parasites. Copeia 1997, 589-592. Bonte, D., Dyck, H. V., Bullock, J. M., Coulon, A., Delgado, M., Gribbs, M., Lillo, F., Marrone, F., Sicilia, A., Castelli, G. and Zava, B. (2005). An invasive Lehouck, V., Matthysen, E., Mustin, K., Saastamoinen, M. et al. (2012). Costs population of Xenopus laevis (Daudin, 1802) in Italy. Herpetozoa 18, 63-64. of dispersal. Biol. Rev. 87, 290-312. Lillo, F., Faraone, F. P. and Valvo, M. L. (2011). Can the introduction of Xenopus Brown, G. P., Shilton, C., Phillips, B. L. and Shine, R. (2007). Invasion, stress, and laevis affect native amphibian populations? Reduction of reproductive occurrence spinal arthritis in cane toads. Proc. Natl. Acad. Sci. USA 104, 17698-17700. in presence of the invasive species. Biol. Inv. 13, 1533-1541. Brown, G. P., Kelehear, C. and Shine, R. (2013). The early toad gets the worm: Lobos, G. and Jaksic, F. M. (2005). The ongoing invasion of African clawed frogs cane toads at an invasion front benefit from higher prey availability. J. Anim. Ecol. (Xenopus laevis) in Chile: causes of concern. Biodivers. Conserv. 14, 429-439. 82, 854-862. Measey, G. J. and Tinsley, R. C. (1997). Mating behavior of Xenopus wittei (Anura: Burton, O. J., Phillips, B. L. and Travis, J. M. J. (2010). Trade-offs and the Pipidae). Copeia 1997, 601-609. evolution of life- histories during range expansion. Ecol. Lett. 13, 1210-1220. Measey, G. J. and Tinsley, R. C. (1998). Feral Xenopus laevis in South Wales. Casterlin, M. E. and Reynolds, W. W. (1980). Diel activity and thermoregulatory Herpetol. J. 8, 23-27. behavior of a fully aquatic frog: Xenopus laevis. Hydrobiologia 75, 189-191. Measey, G. J., Rö dder, D., Green, S. L., Kobayashi, R., Lillo, F., Lobos, G., Clobert, J., Le Galliard, J.-F., Cote, J., Meylan, S. and Massot, M. (2009). Rebelo, R. and Thirion, J.-M. (2012). Ongoing invasions of the African clawed Informed dispersal, heterogeneity in animal dispersal syndromes and the frog, Xenopus laevis: a global review. Biol. Inv. 14, 2255-2270. dynamics of spatially structured populations. Ecol. Lett. 12, 197-209. Miller, K. (1982). Effect of temperature on sprint performance in the frog Xenopus ̈ De Busschere, C., Courant, J., Herrel, A., Rebelo, R., Rodder, D., Measey, G. J. laevis and the salamander Necturus maculosus. Copeia 1982, 695-698. and Backeljau, T. (2016). Unequal contribution of native South African Phillips, B. L., Brown, G. P., Webb, J. K. and Shine, R. (2006). Invasion and the phylogeographic lineages to the invasion of the African clawed frog, Xenopus evolution of speed in toads. Nature 439, 803-803. laevis, in Europe. PeerJ 4, e1659. Phillips, B. L., Brown, G. P., Travis, J. M. J. and Shine, R. (2008). Reid’s paradox Eggert, C. and Fouquet, A. (2006). A preliminary biotelemetric study of a feral revisited: the evolution of dispersal kernels during range expansion. Am. Nat. 172, invasive Xenopus laevis population in France. Alytes 23, 144-149. S34-S48. Excoffier, L. (2009). Patterns of DNA sequence diversity and genetic structure after Phillips, B. L., Brown, G. P. and Shine, R. (2010). Life-history evolution in range- a range expansion: lessons from the infinite-island model. Mol. Ecol. 13, 853-864. shifting populations. Ecology 91, 1617-1627. Excoffier, L. and Ray, N. (2008). Surfing during population expansion promotes Rebelo, R., Amaral, P., Bernardes, M., Oliviera, J., Pinheiro, P. and Leitao,̃ D. genetic revolutions and structuration. Trends Ecol. Evol. 23, 347-351. (2010). Xenopus laevis (Daudin, 1802), a new exotic amphibian in Portugal. Biol. Faraone, F. P., Lillo, F., Giaccone, G. and Valvo, M. L. (2008). The large invasive Inv. 12, 3383-3387. population of Xenopus laevis in Sicily, Italy. Amphib-Reptilia 29, 405-412. Richards, C. T. (2010). Kinematics and hydrodynamics analysis of swimming Fouquet, A. and Measey, G. J. (2006). Plotting the course of an African clawed frog anurans reveals striking inter-specific differences in the mechanism for producing invasion in Western France. Anim. Biol. 56, 95-102. Fronhofer, E. A. and Altermatt, F. (2015). Eco-evolutionary feedbacks during thrust. J. Exp. Biol. 213, 621-634. experimental range expansions. Nat. Commun. 6, 6844. Robert, J., Abramowitz, L., Gantries, J. and Morales, H. D. (2007). Xenopus Gurdon, J. B. and Hopwood, N. (2000). The introduction of Xenopus laevis into laevis: a possible vector of Ranavirus infection? J. Wildl. Dis. 43, 654-652. developmental biology: of empire, pregnancy testing and ribosomal genes. Shine, R., Brown, G. P. and Phillips, B. L. (2011). An evolutionary process that Int. J. Dev. Biol. 44, 43-50. assembles phenotypes through space rather than through time. Proc. Natl. Acad. Hallatschek, O. and Nelson, D. R. (2007). Gene surfing in expanding populations. Sci. USA 108, 5708-5711. Theor. Pop. Biol. 73, 158-170. Stevens, V. M., Pavione, S. and Baguette, M. (2010). Variation within and between Herrel, A., Gonwouo, L. N., Fokam, E. B., Ngundu, W. I. and Bonneaud, C. closely related species uncovers high intra-specific variability in dispersal. PLoS (2012). Intersexual differences in body shape and locomotor performance in the ONE 5, e11123. aquatic frog, Xenopus tropicalis. J. Zool. 287, 311-316. Travis, J. M. J. and Dytham, C. (2002). Dispersal evolution during invasions. Evol. Hudson, C. M., Phillips, B. L., Brown, G. P. and Shine, R. (2015). Virgins in the Ecol. Res. 4, 1119-1129. vanguard: low reproductive frequency in invasion-front cane toads. Biol. J. Linn. Travis, J. M. J., Mü nkemü ller, T. and Burton, O. J. (2010). Mutation and the Soc. 116, 743-747. evolution of dispersal during range expansions. J. Evol. Biol. 23, 2656-2667. Hughes, C. L., Hill, J. K. and Dytham, C. (2003). Evolutionary trade-offs between Videler, J. J. and Jorna, J. T. (1985). Functions of the sliding pelvis in Xenopus reproduction and dispersal in populations at expanding range boundaries. laevis. Copeia 1985, 251-254. Proc. R. Soc. B 270, S147-S150. Wilson, J. R. U., Dormontt, E. E., Prenais, J. P., Lowe, J. A. and Richardson, Karlsson, B. and Johansson, A. (2008). Seasonal polyphenism and M. D. (2009). Something in the way you move: dispersal pathways affect invasion developmental trade-offs between flight ability and egg laying in a pierid success. Trends Ecol. Evol. 24, 136-144. butterfly. Proc. R. Soc. B 275, 2131-2136. Zug, G. R. (1978). Anuran locomotion–structure and function, 2: jumping Kubisch, A., Fronhofer, E. A., Poethke, H. J. and Hovestadt, T. (2013). Kin performance of semiaquatic, terrestrial, and arboreal frogs. Smithson. Contrib. competition as a major driving force for invasions. Am. Nat. 181, 700-706. Zool. 276, 1-31. Journal of Experimental Biology

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APPENDIX I

De Busschere C, Courant J, Herrel A, et al (2016) Unequal contribution of native South African phylogeographic lineages to the invasion of the African clawed frog, Xenopus laevis, in Europe. PeerJ 4:e1659.

Unequal contribution of native South African phylogeographic lineages to the invasion of the African clawed frog, Xenopus laevis, in Europe

Charlotte De Busschere1, Julien Courant2, Anthony Herrel2, Rui Rebelo3, Dennis Rödder4, G. John Measey5 and Thierry Backeljau1,6

1 Operational Direction Taxonomy and Phylogeny, Royal Belgian Institute of Natural Sciences, Brussels, Belgium 2 UMR7179, Département d’Ecologie et de Gestion de la Biodiversité, Centre national de la recherche scientifique, Paris, France 3 Departamento de Biologia Animal/ Centre for Ecology, Evolution and Environmental Changes, Faculdade de Ciências da Universidade de Lisboa, Lisboa, Portugal 4 Herpetology Department, Zoologisches Forschungsmuseum Alexander Koenig, Bonn, Germany 5 Centre of Invasive Biology, Department of Botany and Zoology, University of Stellenbosch, Stellenbosch, South-Africa 6 Evolutionary Ecology Group, University of Antwerp, Antwerp, Belgium

ABSTRACT Due to both deliberate and accidental introductions, invasive African Clawed Frog (Xenopus laevis) populations have become established worldwide. In this study, we investigate the geographic origins of invasive X. laevis populations in France and Portugal using the phylogeographic structure of X. laevis in its native South African range. In total, 80 individuals from the whole area known to be invaded in France and Portugal were analysed for two mitochondrial and three nuclear genes, allowing a comparison with 185 specimens from the native range. Our results show that native phylogeographic lineages have contributed differently to invasive European X. laevis Submitted 19 October 2015 populations. In Portugal, genetic and historical data suggest a single colonization Accepted 13 January 2016 event involving a small number of individuals from the south-western Cape region Published 1 February 2016 in South Africa. In contrast, French invasive X. laevis encompass two distinct native Corresponding author phylogeographic lineages, i.e., one from the south-western Cape region and one from Charlotte De Busschere, the northern regions of South Africa. The French X. laevis population is the first example [email protected] of a X. laevis invasion involving multiple lineages. Moreover, the lack of population Academic editor structure based on nuclear DNA suggests a potential role for admixture within the Joanna Elson invasive French population. Additional Information and Declarations can be found on page 15 Subjects Biodiversity, Biogeography, Ecology, Genetics DOI 10.7717/peerj.1659 Keywords Phylogeography, Population genetics, Invasion history, Invasive species, Xenopus laevis Copyright 2016 De Busschere et al.

Distributed under INTRODUCTION Creative Commons CC-BY 4.0 Reconstructing the invasion history and dynamics of invasive species is crucial for OPEN ACCESS understanding biological invasions and for developing appropriate management strategies

How to cite this article De Busschere et al. (2016), Unequal contribution of native South African phylogeographic lineages to the invasion of the African clawed frog, Xenopus laevis, in Europe. PeerJ 4:e1659; DOI 10.7717/peerj.1659 (Sakai et al., 2001; Lee & Gelembiuk, 2008; Prentis et al., 2009). Moreover, exploring patterns of population genetic variation and evolutionary processes may be key to infer the invasive potential of invasive alien species (Sakai et al., 2001; Lee & Gelembiuk, 2008). In this study, we focus on the African clawed frog, Xenopus laevis (Daudin, 1802), which is indigenous to Southern Africa, including South Africa up to Malawi (i.e., X. laevis sensu stricto Furman et al., 2015). Due to deliberate and accidental introductions from laboratories and pet suppliers, invasive X. laevis populations have become established in Asia, Europe, North America and South America (Tinsley & McCoid, 1996; Lobos & Measey, 2002; Crayon, 2005; Fouquet & Measey, 2006; Faraone et al., 2008; Measey et al., 2012; Peralta-García, Valdez-Villavicencio & Galina-Tessaro, 2014). Invasive X. laevis populations have negative impacts on local biota by reducing the occurrence of reproduction (Lillo, Faraone & Lo Valvo, 2010) and increasing predation pressures on native prey organisms (Lafferty & Page, 1997; Faraone et al., 2008; Measey et al., 2015). Lobos & Measey (2002) also suggested that X. laevis might have indirect impacts on the aquatic system such as increasing water turbidity and nutrient release. Finally, the spread of Batrachochytrium dendrobatis, which causes the amphibian skin disease chytridiomycosis and negatively impacts amphibian populations (Berger et al., 1998; Lips et al., 2006; Skerratt et al., 2007; Voyles et al., 2009; Crawford, Lips & Bermingham, 2010), has been linked to invasive amphibian species such as X. laevis which are often asymptotic carriers (Weldon et al., 2004); however, this link has yet to be proven. Measey et al. (2012) suggested that the global potential invasiveness of X. laevis has been severely underestimated and that it is likely that X. laevis will expand its present colonized area. For example in Europe, introduced X. laevis populations are currently established in France (Fouquet, 2001), Portugal (Rebelo et al., 2010) and Italy (Sicily) (Lillo et al., 2005), though the predicted suitable climate space for X. laevis covers over one million km2 making this a species of European concern (Measey et al., 2012). Within its native range, X. laevis has a wide geographical distribution in which it occupies a variety of natural, as well as manmade waterbodies (Measey, 2004). Native X. laevis populations are distributed from winter rainfall regions in the south-western Cape region to summer rainfall regions in the north; and from sea level to nearly 3,000 m in Lesotho (Measey, 2004). Furthermore, several physiological and behavioural traits enable X. laevis to cope with, dehydration, high levels of salinity, starvation and anoxic conditions (reviewed in Measey et al., 2012). Throughout the native range of X. laevis significant population differentiation has been observed based on both mitochondrial (mtDNA) and nuclear DNA (nDNA) sequences (Grohovaz, Harley & Fabian, 1996; Measey & Channing, 2003; Du Preez et al., 2009; Furman et al., 2015). As such, Furman et al. (2015) identified four phylogeographic lineages within South Africa based on mtDNA and nDNA: (1) south-western Cape (SA1–SA2), (2) Beaufort West (SA4), (3) Niewoudtville (SA7) and (4) northern South Africa (Kimberley, Victoria West, Potchefstroom, SA5) (Fig. 1). The latter two clades are geographically separated by the coastal regions due to the Great Escarpment i.e., a plateau edge running parallel with the South African coast and separating the inland plateau from the coastal plains (Grab, 2010). An admixture zone was observed around Laingsburg (SA3) between south-western Cape (SA1–SA2) and

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 2/21 Beaufort West (SA4). Laingsburg (SA3) and Beaufort West (SA4) are both located within the lowland part of the Great Karoo (700–800 m above sea level) which is separated from the south-western Cape (SA1–SA2) by the Cape Fold Mountains (Du Preez et al., 2009; Furman et al., 2015) and from the north by the Great Escarpment (Fig. 1). Another admixture zone was suggested between Niewoudtville (SA7) and the south-western Cape populations (SA1) around Vredendal (Measey & Channing, 2003). In its native range X. laevis also displays substantial phenotypic population differentiation. Du Preez et al. (2009), for example, showed that male X. laevis from south of the Cape Fold Mountains were longer and heavier than males from north of the Cape Fold Mountains. Given its wide native geographical and ecological ranges and high population genetic diversity, it is important to identify the source areas and population(s) of invasive X. laevis populations if one aims to understand the invasion history and invasive potential of this species (Sakai et al., 2001; Lee, 2002; Dlugosch & Parker, 2008; D’Amen, Zimmermann & Pearman, 2013). The origin of invasive X. laevis populations in Sicily and Chile has already been explored using DNA markers (Lillo et al., 2013; Lobos et al., 2014). Similarly, the origin of specimens from animal suppliers i.e., Xenopus-1 Inc. (Dexter, MI, USA) and Xenopus Express (Brooksville, FL, USA), has been assessed (Du Preez et al., 2009). These studies support the assumption that export of X. laevis specimens for laboratory use mainly stemmed from the south-western Cape region in South Africa (Tinsley & McCoid, 1996; Weldon, De Villiers & Du Preez, 2007), although other source areas cannot be excluded, especially for older introduced stocks (Du Preez et al., 2009; L Van Sittert & J Measey, 2015, unpublished data). The present paper aims to unravel to which extent single or multiple native phylogeographic lineages have contributed to the invasive populations in France and Portugal by comparing mtDNA and nDNA data sampled across the invaded, as well as the native range. In France, it is assumed that X. laevis was introduced at Bouillé-Saint-Paul (Deux Sèvres) from a nearby breeding facility where X. laevis was bred from the 1950s until 1996 (Fouquet, 2001; Fouquet & Measey, 2006). From that breeding facility X. laevis may have escaped repeatedly and was probably released when the facility was definitively closed in 1996 (Measey et al., 2012). Currently, French X. laevis populations occupy an area of approximately 200 km2 near the city of Saumur (Maine-et-Loire) (Fig. 1). The introduction of X. laevis in Portugal is assumed to be accidental, caused in 1979 by the inundation of a basement of a research institute at Oeiras along river Laje, about 20 km west of Lisbon (Rebelo et al., 2010; Measey et al., 2012). Nowadays, several populations are found in two tributaries of river Tagus, i.e., river Laje and river Barcarena (Fig. 1) (Rebelo et al., 2010).

METHODS Taxon sampling In total 80 individuals from 32 localities, covering the known area invaded by X. laevis in France (FR) and Portugal (PT) were captured. For comparison with previous work of Lillo et al. (2013) we included two specimens from Sicily (IT, provided by Francesco

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 3/21 Figure 1 Map of the native (A. South Africa) and invaded X. laevis localities (B. Portugal, C. France) surveyed in this study. Abbreviations and colours of sampling localities (circles) refer to geographical regions that are mentioned in methods (see ‘Taxon sampling’). More detailed locality information is provided in Online Resource 1. National and provincial borders of South African Provinces are visualized by solid and dashed lines respectively (A). Rivers and roads are represented by blue and grey lines respectively (B) and (C). Names of main rivers (italic) and towns are shown (B) and (C).

Lillo). Within the native range of X. laevis, 21 specimens were sampled. Seven specimens came from two localities in the south-western Cape (SA1) i.e., Cape of Good Hope nature reserve and a single dam at Jonkershoek, the location from which animals were shipped for international trade from the 1940s to 1970s (L Van Sittert & J Measey, 2015, unpublished data). 14 specimens came from five localities around Rooikrantz Dam near the historical Pirie hatchery nearby King Williams Town, Eastern Cape (SA6) (L Van Sittert

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 4/21 & J Measey, 2015, unpublished data). These new data were supplemented by data on 164 South African specimens available on GenBank. One individual of Xenopus gilli Rose & Hewitt, 1927 was sampled near Kleinmond (South Africa). Based on Furman et al. (2015), sampling localities in South Africa were grouped into seven geographical regions: SA1: south-western Cape, south-west of the Cape Fold Mountains up to De Doorns; SA2: Cape, Hoekwil & Tsitsikamma region; SA3: Cape, Laingsburg; SA4: Cape, Beaufort West; SA5: northern South Africa (Kimberley, Victoria West, Potchefstroom); SA6: Cape, Rooikrantz Dam; SA7: Nieuwoudtville (Fig. 1). All specimen data with full locality information are provided in Online Resource 1. Animals from the Portuguese invasive population were captured under the permit no 570/2014/CAPT from Instituto da Conservação da Natureza e das Florestas, in the scope of the ‘‘Plano de erradicação de Xenopus laevis nas ribeiras do Concelho de Oeiras’’. Animals from the native South-African populations were sampled under the permits 0056-AAA007-00092 (CapeNature, SA1) and CRO 109/13CR (Eastern Cape, SA6) provided by the Department of Economic Development, Environmental Affairs and Tourism and with ethics approval from the Research Ethics Committee: Animal Care and Use (protocol number: SU-ACUD14-00028). DNA ampliﬁcation and sequencing Invasive specimens were euthanized with a lethal injection of sodium pentobarbital. Muscle tissue was dissected from invasive specimens, while native wild caught specimens were toe-clipped to obtain tissue samples. Genomic DNA was extracted by means of a NucleoSpin R tissue kit (Macherey-Nagel, Düren, Germany) according to the manufacturer’s protocol. Five genomic DNA fragments (circa (ca.) 2,040 bp) representing two mitochondrial and three nuclear gene fragments were amplified using PCR. These genes were selected in order to enable comparison with previously published work on native (Measey & Channing, 2003; Du Preez et al., 2009; Bewick, Anderson & Evans, 2011; Furman et al., 2015) and invasive X. laevis specimens (Lillo et al., 2013; Lobos et al., 2014). Fragments of the mitochondrial cytochrome b gene (Cytb; ca. 280 bp) and 16S ribosomal DNA (16S; ca. 800 bp) were amplified and sequenced with the primer pairs Cytb I /Cytb II and 16Sc-L/16Sd-H (Kessing et al., 1989; Evans et al., 2003). Fragments of the nuclear protein coding genes arginine methyltransferase 6 (Prmt6; ca. 666 bp), androgen receptor isoform α (AR; ca. 402 bp) and microtubule associated serine/threonine kinase-like protein (Mastl; ca. 539 bp) were amplified and sequenced with the primer pairs Exon4_for1/Exon4_rev2, XLAR_for_40/XLAR_rev_431 and Exon13_fora/Exon13_reva (Bewick, Anderson & Evans, 2011). The nuclear primer pairs are assumed to be paralog-specific, hence only amplifying one pair of alleles (see Bewick, Anderson & Evans, 2011). PCR amplifications were run with the conditions reported in Online Resource 3. PCR products were purified with FastAP TM thermosensitive alkaline phosphatase in combination with exonuclease I and subsequently sequenced in both directions. Nucleotide sequences were assembled and edited in CodonCode Aligner (CodonCode Corporation, Dedham, MA, USA). They were aligned together with corresponding well-documented sequences of South African X. laevis available from GenBank using the ClustalW algorithm (Thompson, Higgins & Gibson, 1994) in MEGA v. 6 (Tamura et al., 2013). For comparison, the two Cytb haplotypes found by

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 5/21 Lobos et al. (2014) in Chilean (CL) invasive X. laevis populations were included in the Cytb alignment. Cytb and 16S gene fragments obtained by Lillo et al. (2013) were not included in the current alignments as they only partially overlapped due to the use of different primers. Nuclear sequences were converted into haplotypes using the PHASE algorithm (Stephens, Smith & Donnelly, 2001; Stephens & Donnelly, 2003) implemented in DnaSP v. 5 (Librado & Rozas, 2009). All new sequence data were deposited in Genbank (accession numbers in Online Resource 1: https://figshare.com/s/44b71de473fc11e5829d06ec4b8d1f61). Genetic diversity The following indices of genetic diversity were estimated using DnaSP v. 5 (Librado & Rozas, 2009): numbers of different haplotypes (h), the number of segregating sites (S), nucleotide diversity (π, i.e., the average number of nucleotide differences per site between two sequences) and haplotype/allelic diversity (Hd, i.e., the probability that two haplotypes drawn uniformly at random from a population are not the same). mtDNA analysis Phylogenetic relationships among invasive and native X. laevis mtDNA sequences were reconstructed with Bayesian inference (BI) and Maximum Parsimony (MP). Three alignments were created: two separate Cytb and 16S alignments including only unique alleles and a concatenated Cytb-16S alignment including only unique haplotypes. In all analyses, X. gilli was used as outgroup. Parsimony informative sites were calculated with DnaSP v. 5 (Librado & Rozas, 2009). BI analyses were performed using MrBayes v. 3.2.4 (Ronquist & Huelsenbeck, 2003). Cytb, 16S and Cytb-16S alignments were analysed under a general time reversible (GTR) model with all model parameters estimated from the data and a proportion of invariant sites (+I) as selected by jModeltest v. 2.1.7 (Posada, 2008). BI analyses were run with two different Metropolis-coupled Markov chains for 10 million generations with sampling every 1,000th generation. The average standard deviations of split frequencies, the potential scale reduction factors and the plots of likelihood versus generation were evaluated to ensure convergence based on the log file and using Tracer v1.6 (Rambaut et al., 2013). A total of 25% of the trees were discarded as burn-in and posterior probabilities were calculated for each split from the remaining set of trees. MP trees were estimated with a neighbour joining tree as starting tree using the Phangorn package (Schliep, 2011) in R 3.1.1 (R Development Core Team, 2014). A set of most-parsimonious trees was generated using the parsimony ratchet (Nixon, 1999) with nearest neighbour interchange rearrangement, 10,000 ratchet iterations and up to maximum 10 rounds. Parsimony bootstrap values were obtained with 1,000 bootstrap replicates using the bootstrap.phyDat function in R 3.1.1. All topologies were visualized and edited in respectively Figtree v. 1.4.2 (Rambaut, 2014) and TreeGraph2 v. 2.4.0-456 beta (Stöver & Müller, 2010). Differentiation between native geographical areas and invaded regions was quantified using pairwise Fst values computed based upon nucleotide pairwise distances and tested for significance with 999 permutations using Arlequin v. 3.5.3.1 (Excoffier & Lischer, 2010). The population comparison was done solely for the 16S alignment as this dataset included sequences of most individuals (n = 179) and representatives of all populations.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 6/21 Autosomal analysis The minimum number of recombination events within alleles were estimated in DnaSP v.5 (Hudson & Kaplan, 1985; Librado & Rozas, 2009). Given putative recombination, median- joining networks (MJN) (Bandelt, Forster & Rohl, 1999) were constructed in PopART v.1.7.2 (PopART, 2015) for each nuclear locus separately to illustrate the mutational differences between alleles and their distribution among invasive and native populations (Mardulyn, 2012). Pairwise Fst values were computed based on the allele frequencies to show to what extent alleles are differently sorted among populations (e.g., where the same alleles are found in different populations but with different frequencies). Native populations were defined based upon the geographical regions mentioned in ‘Taxon sampling.’ Individuals from France and Portugal respectively were treated as two populations. Pairwise Fst values were computed from the allele frequencies with 999 permutations to generate the probability that a random value would be greater than or equal to the observed data using GenAlEx v. 6.5 (Peakall & Smouse, 2012). Individuals with missing data for more than one gene were excluded from the analysis. In order to visualize population differentiation, the pairwise Fst matrix was subsequently used as a distance matrix for a principal coordinates analysis (PCoA) in GenAlEx v.6.5.

RESULTS mtDNA Mitochondrial alignments were constructed for Cytb and 16S involving 101 and 181 individuals respectively. The concatenated Cytb-16S alignment involved 64 individuals. None of the alignments showed gaps. The Cytb alignment involved16 unique alleles of X. laevis with 42 variable positions of which 34 were parsimony informative. No Cytb sequences were obtained of specimens from the regions SA2, SA3 and SA4. In contrast, all geographic regions were represented in the 16S alignment, which involved 15 alleles of X. laevis with 25 variable positions of which 17 were parsimony informative. The concatenated Cytb-16S alignment involved seven unique X. laevis haplotypes with 43 variable positions of which 39 were parsimony informative (Online Resource 2: https://figshare.com/s/44b71de473fc11e5829d06ec4b8d1f61). The concatenated alignment comprised only specimens from the invaded regions PT, IT and FR and the native regions SA1 and SA6. The BI and MP trees for Cytb and 16S are shown in Fig. 2. These analyses all resolved consistently three well-supported clades comprising X. laevis alleles from the geographic regions (1) SA5, SA6 and FR, (2) SA7 and (3) SA1, PT, FR, IT. Additionally, the phylogeny generated from the 16S data strongly supported (1) SA1, SA2, SA3, FR, PT and IT and (2) SA3-SA4 as well-supported X. laevis clades. BI and MP trees for the concatenated data (Online Resource 2) showed two highly supported clades comprising haplotypes from (1) FR, PT, IT, SA1 and (2) FR and SA6. Pairwise population Fst values among all geographic regions based on 16S were significant except among (1) SA1, SA2 and PT, (2) SA5 and SA6 and among (3) FR and SA6 (Online Resource 4).

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 7/21 Figure 2 MP and BI inference based on Cytb (A) and 16S alignments (B). Bayesian consensus trees are visualized with Posterior BI bootstrap (B > 0.70) and Parsimony bootstrap values. Parsimony scores (i.e., tree length) of Cytb MP tree and 16S MP tree were 83 and 47 respectively. Geographical regions where alleles have been observed are indicated in grey (abbreviations see Fig. 1 and ‘Methods’).

nDNA AR sequences (n = 121) showed eight polymorphic positions of which seven revealed heterozygous genotypes. The maximum number of heterozygous positions observed within a single AR sequence was three but that was only observed in two individuals. The two most frequent AR alleles i.e., hap_1 and hap_2 differed by two nonsynonymous substitutions. The Mastl alignment (n = 180) showed 37 polymorphic positions and 108 individuals were heterozygous at 3 ± 2 positions on average. The Prmt6 alignment (n = 176) showed 20 polymorphic positions and 107 individuals were heterozygous at on average 2 ± 1 positions. MJN revealed high allelic diversity at Mast1 and Prmt6 and low allelic diversity at AR (Table 1). In the MJN, the most frequent alleles were often found in different geographical regions (Fig. 4). When combining the allelic frequency information of the three nDNA genes, pairwise population Fst values among all geographic regions were significant except between SA5 and SA6 (Online Resource 5: https://figshare.com/s/44b71de473fc11e5829d06ec4b8d1f61). The first two PCoA-axes explained 75% of the total variance from the pairwise Fst values. There was a clear separation among SA7 and the remnant populations along the first PCoA axis (explaining 28% of the total variance). Negative values along the second PCoA axis (explaining 47% of the total variance) were linked to the northernmost up to the north-eastern native populations (SA4–SA7), while positive values were linked to the invasive (Fr, PT) and southernmost

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 8/21 Table 1 Summary of diversity statistics for all loci and each population. Statistics describing the number of nucleotides sequenced (nt), number of individuals (ind), number of different haplotypes (h), number of segregating sites (S), nucleotide diversity (π) and haplotype diversity (Hd).

Gene Country Pop nt ind h S π Hd Cytb France FR 282 42 2 26 0.0233 0.251 Portugal PT 282 18 2 1 0.00105 0.294 South Africa SA1 282 12 3 3 0.00265 0.53 SA5 282 13 8 10 0.01099 0.91 SA6 282 6 2 1 0.00119 0.333 SA7 282 4 2 6 0.01429 0.667 16S France FR 544 56 2 13 0.00598 0.249 Portugal PT 544 16 1 0 0 0 South Africa SA1 544 23 2 1 0.00055 0.3 SA2 544 14 2 1 0.00026 0.143 SA3 544 9 3 11 0.00756 0.556 SA4 544 12 3 2 0.00106 0.53 SA5 544 13 3 2 0.0008 0.41 SA6 544 14 3 3 0.00079 0.275 SA7 544 22 2 1 0.00068 0.368 AR France FR 294 58 2 2 0.00338 0.495 Portugal PT 294 18 1 0 0 0 South Africa SA1 294 27 1 0 0 0 SA2 294 13 4 4 0.0044 0.582 SA3 294 18 2 2 0.00347 0.508 SA4 294 20 3 2 0.0014 0.229 SA5 294 25 3 2 0.00078 0.222 SA6 294 9 1 0 0 0 SA7 294 23 2 1 0.00373 0.043 Mastl France FR 525 53 13 6 0.00499 0.804 Portugal PT 525 18 2 5 0.00053 0.056 South Africa SA1 525 27 10 10 0.00609 0.73 SA2 525 12 9 8 0.00662 0.873 SA3 525 9 8 12 0.00712 0.895 SA4 525 13 8 9 0.00453 0.745 SA5 525 13 13 10 0.00452 0.938 SA6 525 9 13 6 0.00421 0.961 SA7 525 24 7 10 0.00264 0.543 Prmt6 France FR 396 51 16 8 0.00679 0.839 Portugal PT 396 18 4 3 0.00213 0.605 South Africa SA1 396 27 10 7 0.00375 0.829 SA2 396 13 9 6 0.00447 0.871 SA3 396 8 10 9 0.00787 0.942 SA4 396 12 7 7 0.00498 0.79 SA5 396 13 13 8 0.00679 0.911 SA6 396 10 15 13 0.00793 0.974 SA7 396 22 1 0 0 0

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 9/21 Figure 3 Result of Principal Co-ordinate analysis of nuclear genetic variation among native (SA1– SA7) and invasive French (FR) and Portuguese (PT) X. laevis populations. PCoA of pairwise Fst values based on allele frequencies in three nuclear loci (n = 180 individuals; Online Resource 5). Abbreviations refer to geographical regions (see Fig. 1).

native populations (SA1–SA3)(Fig. 3). Native populations SA5 and SA6 clustered together (Fig. 3). South Africa Haplotype diversity of nDNA and mtDNA within the native range of X. laevis was on average 0.60 ± 0.20 and 0.45 ± 0.18 respectively. Mitochondrial nucleotide diversity within the native regions was either rather low (π ≤ 0.0016: SA1, SA2, SA4, SA6) or high (π ≥ 0.0059: SA3, SA5, SA7). As mentioned previously, mtDNA and nDNA variation was geographically structured within South Africa (Fig. 2, Online Resource 4 and 5) (Furman et al., 2015). The monophyly of the sequences from northern South African regions (SA5) and the newly sampled sites at Rooikrantz Dam in the Eastern Cape (SA6) was consistently strongly supported by both Cytb and 16S (Fig. 2). Furthermore, the pairwise population Fst values based on nDNA and on 16S separately were not significant between these two regions (Online Resource 4 and 5: https://figshare.com/s/44b71de473fc11e5829d06ec4b8d1f61). Portugal Nucleotide and haplotype diversities in Portugal were extremely low for all markers (0 ≤ π < 0.00105, 0 ≤ Hd < 0.294, Table 1), except for Prmt6 (π = 0.00213, Hd = 0.605, Table 1). Only two alleles were exclusively found in Portugal, viz. one for Cytb (hap_11) and one for Prmt6 (hap_27). The concatenated Cytb-16S haplotype (hap_11_2) was only found in Portugal. All other mtDNA and nDNA alleles occurred also within native South African populations. More precisely, identical mtDNA was found in X. laevis populations from the south-western Cape SA1 (Cytb: hap_12) and from the Cape regions SA1–SA3 (16S: hap_2). Identical AR, Prmt6 and Mastl alleles were found in respectively the following native regions (1) SA1–SA3, (2) SA1–SA4 and SA7 and (3) SA1–SA7. Pairwise Fst values based on 16S were not significant among Portugal and native regions SA1 and SA2 (Online Resource 4). PCoA and MJN revealed that nuclear allele frequencies within Portugal were most similar to those of native region SA1 (Fig. 3, Online Resource 5). The mean Portuguese nucleotide diversity (π = 0.0007 ± 0.0009) across all loci was much lower than

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 10/21 Figure 4 MJN of nuclear Prmt6 (A), AR (B) and Mastl (C) sequence data from native and invasive X. laevis populations. The sizes of the circles are proportional to allele frequencies. Colours refer to native geographic and invaded regions (see legend). Small black nodes represent unsampled alleles and numbers of mutations are marked by stripes on the connecting branches.

in native populations (π = 0.0039 ± 0.0035). Similarly, the mean Portuguese haplotype diversity (Hd = 0.191 ± 0.261) across all loci was lower than in the native populations (Hd = 0.558 ± 0.323). France Two distinct alleles were observed in the introduced French population for both Cytb and 16S (Table 1, Fig. 2). Moreover, these alleles were highly similar or identical to alleles found in two geographically non-overlapping native regions in South Africa, i.e., the northern South Africa and Rooikrantz Dam region (SA5–SA6) and the south-western Cape region (SA1–SA3) (Fig. 2). The most abundant Cytb allele in France (i.e., hap_16;

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 11/21 86% of specimens) was very similar to Cytb sequences in northern South Africa and Rooikrantz Dam populations (SA5–SA6) with on average four nucleotide differences (Fig. 2). Similarly, 86% of the French individuals had a 16S allele (i.e., hap_12) identical to a 16S sequence only found in northern South Africa and Rooikrantz Dam region (SA5–SA6)(Fig. 2). The concatenated Cytb-16S dataset comprised 38 French individuals which represented two different haplotypes. One haplotype was identical to a haplotype found in SA1 (i.e., hap_12_2) and the other haplotype was only found in France but highly similar to haplotypes found in SA6 (i.e., hap_16_12; Online Resource 2). MtDNA haplotype diversity for each marker was low within the French population (Hd ≤ 0.251, Table 1). Pairwise Fst value based on 16S among SA6 and France was not significant (Fst = 0.049; Online Resource 4). Haplotype diversities for nDNA ranged from 0.50 for AR up to 0.80 for Mastl and Prmt6 (Table 1). France comprised 13 Mastl and 16 Prmt6 alleles with seven Mastl and 10 Prmt6 alleles being hitherto only found in France. nDNA allele frequencies in France were most similar to those in native populations from Laingsburg in the south-western Western Cape Province (SA3, Fig. 3, Online Resource 4: https://figshare.com/s/44b71de473fc11e5829d06ec4b8d1f61). No significant difference was observed when comparing nDNA allele frequencies among individuals representing the two different mitochondrial groups (Fst = 0; p = 0.421), however sample sizes were low i.e., 49 and 8 individuals representing northern South Africa-Rooikrantz Dam mtDNA (SA5–SA6) and south-western Cape mtDNA (SA1–SA2) respectively. In comparison with the native geographic regions, mean nucleotide diversity across all loci in France was higher (π = 0.009 ± 0.008 versus π = 0.0039 ± 0.0035) while French haplotype diversity across all loci was comparable (Hd = 0.528 ± 0.287 versus Hd = 0.558 ± 0.323). Sicily The Sicilian individuals were identical to native individuals from SA1 for Cytb and from SA1–SA3 for 16S (Fig. 1). Only one concatenated Cytb-16S haplotype was found which was unique to Sicily and highly similar to haplotypes found in FR, PT and SA1. Sicilian AR and Prmt6 alleles were identical to alleles occurring in the native range regions SA1-SA3. Cytb and 16S sequences of Sicilian individuals sampled by Lillo et al. (2013) were not included in the current analyses as they involved different gene regions. However, the mtDNA sequences in the present study were identical to those of Lillo et al. (2013) in the overlapping gene regions (Cytb: ∼242 bp, 16S: ∼374 bp).

DISCUSSION Since the 1930s, X. laevis has successfully invaded extensive areas worldwide particularly due to its popularity for laboratory use and pet trade (Tinsley, Loumont & Kobel, 1996; Gurdon & Hopwood, 2000; Lobos & Measey, 2002; Crayon, 2005; Faraone et al., 2008; Measey et al., 2012; Herrel & Van Der Meijden, 2014). Reconstructing the invasion history of X. laevis is pivotal to understand the invasion biology and range dynamics of this species and as such, might be critical for developing future management strategies (Sakai et al., 2001; Lee & Gelembiuk, 2008; Prentis et al., 2009). The identification of source populations is particularly relevant when there is extensive population differentiation within the

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 12/21 native range because, as well phenotypic as genotypic traits of colonizing individuals might influence the invasion process (Sakai et al., 2001; Lee & Gelembiuk, 2008). Here, the origin of invasive French and Portuguese X. laevis specimens was investigated using DNA sequence data. The genetic diversity of the invasive Portuguese populations was lower than across the native range of X. laevis in South Africa, but the Portuguese sequences were very similar or even identical to those of native individuals from the south-western Cape region in South Africa. This suggest (1) that the Portuguese populations may be derived from the latter and (2) that the Portuguese population derives from a single colonization event involving a small number of individuals, most likely stemming from one and the same source population in the south-western Cape region. This is in line with what one would expect from the historical data which attributes the introduction of X. laevis in Oeiras (Portugal) to a single accidental flood of a basement of a research institute in 1979 (Rebelo et al., 2010). Analogous to the Portuguese samples, the identical sequences shared by the Sicilian (Lillo et al., 2013), Chilean (Lobos et al., 2014) and south-western Cape populations samples support the idea that X. laevis was imported in Sicily and Chile from wild populations in the south-western Western Cape Province. Indeed, the export of X. laevis from the Western Cape Province is well documented, especially between 1940 and 1974 when there was only one official supplier in South Africa, i.e., Jonkershoek Fish Hatchery (Weldon, De Villiers & Du Preez, 2007). In contrast to the X. laevis colonization in Portugal, two distinct and divergent mtDNA lineages were detected in France. These lineages were related to two geographically non- overlapping native regions in South Africa. A majority of the French individuals possessed mtDNA highly similar or identical to a phylogeographic lineage from northern South Africa and Rooikrantz Dam (SA5–SA6, further referred to as the northern lineage). The other French individuals had mtDNA identical to native individuals from the south-western Cape (SA1–SA2). Haplotype and sequence diversity in the French population were relatively high and comparable to those in native South African regions. The mtDNA data thus suggest that two distinct phylogeographic lineages i.e., south-western Cape and northern lineage, contributed to the invasion of X. laevis in France. Although the nDNA data are consistent with this suggestion, they solely indicate the Laingsburg region (SA3) and the south- western Cape (SA1–SA2) as possible source areas of the French nDNA alleles. This might be explained by either a scenario of a single South African source area where south-western and northern lineages admix or a scenario in which France was invaded by individuals from several distinct South African source areas followed by admixture of the colonizing animals. In the case of the first scenario, the most likely source population based upon the nDNA data would be the Laingsburg region which is considered as an admixture zone among the south-western Cape regions SA1-SA2 and the Beaufort West region SA4 (Du Preez et al., 2009; Furman et al., 2015). However, individuals from Laingsburg do not show mtDNA of the northern populations. Hence, the occurrence of the main mitochondrial alleles (Cytb: hap_16; 16S: hap_12) in French individuals cannot be explained by this scenario. Conversely, historical data supports the second scenario in which the X. laevis invasion in France stems from different source areas as animal suppliers are known from

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 13/21 both the south-western Western Cape province and the Eastern Cape Province (Port Elizabeth)(L Van Sittert & J Measey, 2015, unpublished data). Unfortunately, information concerning export from the latter region is limited. In contrast, there is some export data from the Western Cape province in the period that the French breeding facility was active (1950–1996) (Weldon, De Villiers & Du Preez, 2007). Until 1974, there was one official animal supplier i.e., Jonkershoek Fish Hatchery. Subsequently, trading was left to private enterprises for which export and collection information is hitherto unknown. Yet, since 1990, permits were licenced to four South African animal suppliers restricting the collection of X. laevis to man-made water bodies (Weldon, De Villiers & Du Preez, 2007). Around the same time, in 1989, the ownership of the French breeding facility changed (Measey et al., 2012). Taking all these circumstances together, it seems likely that during its 56 years of existence, the French breeding facility might have imported X. laevis repeatedly from different South African sources. Moreover, it seems not unlikely that secondary trading occurred within Europe or even worldwide e.g., among commercial breeding facilities and/or research institutes. To the best of our knowledge, the French X. laevis population is the first example in which two geographically non-overlapping phylogenetic lineages participated in a X. laevis invasion. As mentioned previously, two distinct mtDNA lineages, related to on the one hand the northern regions of South Africa and on the other hand the southwestern Cape regions, are present in the French population. However, no population structure could be found in the invasive French populations based on nDNA. In contrast, within the native South African range these mtDNA lineages are significantly divergent based on nDNA. The latter observation suggests a potential role for ongoing admixture within the invasive French population. Likewise, admixture might have occurred within breeding facilities among specimens from different source populations preceding the French invasion. The combination of genetic variation from multiple phylogeographic lineages might explain the high level of genetic diversity within France, which was comparable to, or even higher than, the diversity observed in native regions. An increase in genetic diversity within invasive population relative to native populations due to multiple introductions of genetically divergent source populations has also been demonstrated in invasions of other organisms elsewhere (Novak & Mack, 1993; Kolbe et al., 2004; Kolbe et al., 2008; Lavergne & Molofsky, 2007). Moreover, intraspecific admixture among previously isolated multiple divergent genetic lineages is often suggested to play an important role in driving the success of colonising populations (Sakai et al., 2001; Lavergne & Molofsky, 2007; Kolbe et al., 2008; Lee & Gelembiuk, 2008; Rius & Darling, 2014). Yet, a species’ invasiveness is not only function of the amount of genetic variation, but even more importantly of the nature of adaptive genetic variation (Dlugosch et al., 2015). Concerning X. laevis, it is clear that single introduction events even with low levels of genetic diversity have proven to be highly successful in invading non-native areas such as Chile and Sicily. However, in order to predict the species’ potential range expansion the intraspecific variation should be taking into account (Pearman et al., 2010; D’Amen, Zimmermann & Pearman, 2013), especially for the French population, in which two genetically distinct and environmentally divergent South African phylogeographic lineages have contributed to the invasion. Hence,

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 14/21 the French population offers a study system for investigating the extent to which the combination of genetic variation from divergent phylogeographic lineages might have influenced the French invasion and/or might influence its future range expansion. In sum, the genetic structure of X. laevis in its native South African range allowed us to investigate the geographic origins of invasive X. laevis populations in France and Portugal. The current analyses showed that native phylogeographic lineages are not equally represented in invasive European X. laevis populations.

ACKNOWLEDGEMENTS This work was conducted in the contexts of the Biodiversa project BR/132/A1/INVAXEN- BE ‘‘Invasive biology of Xenopus laevis in Europe: ecology, impact and predictive models’’ and of the BELSPO-IUAP project P7/04 ‘‘SPEEDY: SPatial and environmental determinants of Eco-Evolutionary DYnamics: anthropogenic environments as a model’’ and FWO research community W0.009.11N ‘‘Belgian Network for DNA Barcoding,’’ coordinated by the ‘‘Joint Experimental Molecular Unit (JEMU)’’ at RBINS. JM would like to thank the DST-NRF Centre of Excellence for Invasion Biology for support. We gratefully thank Francesco Lillo, François Lefebvre, Jean Secondi, Mohlamatsane Mokhatla and Lubabalo Mofu for their aid in sampling Xenopus specimens.

ADDITIONAL INFORMATION AND DECLARATIONS

Funding This work was financially supported by the Biodiversa project BR/132/A1/INVAXEN-BE ‘‘Invasive biology of Xenopus laevis in Europe: ecology, impact and predictive models’’ at RBINS. JM received funds from the National Research Foundation (South Africa) CPRR13080726510 and incentive funding. The funders had no role in study design, data collection and analysis, decision to publish, or preparation of the manuscript. Grant Disclosures The following grant information was disclosed by the authors: Biodiversa project: BR/132/A1/INVAXEN-BE. National Research Foundation (South Africa): CPRR13080726510. Competing Interests John Measey is an Academic Editor for PeerJ. Author Contributions • Charlotte De Busschere conceived and designed the experiments, performed the experiments, analyzed the data, wrote the paper, prepared figures and/or tables. • Julien Courant, Anthony Herrel, Rui Rebelo and G. John Measey contributed reagents/materials/analysis tools, reviewed drafts of the paper. • Dennis Rödder reviewed drafts of the paper. • Thierry Backeljau conceived and designed the experiments, contributed reagents/materials/analysis tools, wrote the paper, reviewed drafts of the paper.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 15/21 Animal Ethics The following information was supplied relating to ethical approvals (i.e., approving body and any reference numbers): Animals from the Portuguese invasive population were captured under the permit no 570/2014/CAPT from Instituto da Conservação da Natureza e das Florestas, in the scope of the ‘‘Plano de erradicação de Xenopus laevis nas ribeiras do Concelho de Oeiras.’’ Animals from the native South-African populations were sampled under the permits 0056- AAA007-00092 (CapeNature, SA1) and CRO 109/13CR (Eastern Cape, SA6) provided by the Department of Economic Development, Environmental Affairs and Tourism and with ethics approval from the Research Ethics Committee: Animal Care and Use (protocol number: SU-ACUD14-00028). DNA Deposition The following information was supplied regarding the deposition of DNA sequences: New sequences are available from Genbank with accession numbers from KT586615 to KT587048. Data Availability The following information was supplied regarding data availability: Data and online resource can be found at: https://figshare.com/s/44b71de473fc11e5829d06ec4b8d1f61. http://dx.doi.org/10.6084/m9.figshare.1577557.

REFERENCES Bandelt HJ, Forster P, Rohl A. 1999. Median-joining networks for inferring intraspecific phylogenies. Molecular Biology and Evolution 16:37–48 DOI 10.1093/oxfordjournals.molbev.a026036. Berger L, Speare R, Daszak P, Green DE, Cunningham AA, Goggin CL, Slocombe R, Ragan MA, Hyatt AD, McDonald KR, Hines HB, Lips KR, Marantelli G, Parkes H. 1998. Chytridiomycosis causes amphibian mortality associated with population declines in the rain forests of Australia and Central America. Proceedings of the National Academy of Sciences of the United States of America 95:9031–9036 DOI 10.1073/pnas.95.15.9031. Bewick AJ, Anderson DW, Evans BJ. 2011. Evolution of the closely related, sex-related genes DM-W and DMRT1 in African clawed frogs (Xenopus). Evolution 65:698–712 DOI 10.1111/j.1558-5646.2010.01163.x. Crawford AJ, Lips KR, Bermingham E. 2010. Epidemic disease decimates amphibian abundance, species diversity, and evolutionary history in the highlands of central Panama. Proceedings of the National Academy of Sciences of the United States of America 107:13777–13782 DOI 10.1073/pnas.0914115107. Crayon J. 2005. Species account: Xenopus laevis. In: Lannoo MJ, ed. Amphibian declines: the conservation status of United States species. Berkeley: University of California Press, 522–525.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 16/21 D’Amen M, Zimmermann NE, Pearman PB. 2013. Conservation of phylogeographic lineages under climate change. Global Ecology and Biogeography 22:93–104 DOI 10.1111/j.1466-8238.2012.00774.x. Daudin FM. 1802. Histoire naturelle des rainettes, des grenouilles et des crapauds. Paris: Imprimerie de Bertrandet: Chez Levrault. Dlugosch KM, Anderson SR, Braasch J, Cang FA, Gillette HD. 2015. The devil is in the details: genetic variation in introduced populations and its contributions to invasion. Molecular Ecology 24:2095–2111 DOI 10.1111/mec.13183. Dlugosch KM, Parker IM. 2008. Founding events in species invasions: genetic variation, adaptive evolution, and the role of multiple introductions. Molecular Ecology 17:431–449 DOI 10.1111/j.1365-294X.2007.03538.x. Du Preez LH, Kunene N, Hanner R, Giesy JP, Solomon KR, Hosmer A, Van Der Kraak GJ. 2009. Population-specific incidence of testicular ovarian follicles in Xenopus laevis from South Africa: A potential issue in endocrine testing. Aquatic Toxicology 95:10–16 DOI 10.1016/j.aquatox.2009.07.018. Evans BJ, Brown RM, McGuire JA, Supriatna J, Andayani N, Diesmos A, Iskandar D, Melnick DJ, Cannatella DC. 2003. Phylogenetics of fanged frogs: testing bio- geographical hypotheses at the interface of the Asian and Australian faunal zones. Systematic Biology 52:794–819 DOI 10.1080/10635150390251063. Excoffier L, Lischer HEL. 2010. Arlequin suite ver 3.5: a new series of programs to perform population genetics analyses under Linux and Windows. Molecular Ecology Resources 10:564–567 DOI 10.1111/j.1755-0998.2010.02847.x. Faraone FP, Lillo F, Giacalone G, Lo Valvo M. 2008. The large invasive population of Xenopus laevis in Sicily, Italy. Amphibia-Reptilia 29:405–412 DOI 10.1163/156853808785112075. Fouquet A. 2001. Des clandestins aquatiques. Zamenis 6:10–11. Fouquet A, Measey GJ. 2006. Plotting the course of an African clawed frog invasion in Western France. Animal Biology 56:95–102 DOI 10.1163/157075606775904722. Furman BLS, Bewick AJ, Harrison TL, Greenbaum E, Gvoždík V, Kusamba C, Evans BJ. 2015. Pan-African phylogeography of a model organism, the African clawed frog Xenopus laevis. Molecular Ecology 24:909–925 DOI 10.1111/mec.13076. Grab S. 2010. Drakensberg Escarpment: mountains of geomorphic diversity. In: Migon P, ed. Geomorphological landscapes of the world. Dordrecht Heidelberg London New York: Springer, 133–142. Grohovaz GS, Harley E, Fabian B. 1996. Significant mitochondrial DNA sequence divergence in natural populations of Xenopus laevis (Pipidae) from Southern Africa. Herpetologica 52:247–253. Gurdon JB, Hopwood N. 2000. The introduction of Xenopus laevis into developmental biology: of empire, pregnancy testing and ribosomal genes. International Journal of Developmental Biology 44:43–50. Herrel A, Van Der Meijden A. 2014. An analysis of the live reptile and amphibian trade in the USA compared to the global trade in endangered species. Herpetological Journal 24:103–110.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 17/21 Hudson RR, Kaplan NL. 1985. Statistical properties of the number of recombination events in the history of a sample of DNA sequences. Genetics 111:147–164. Kessing BD, Croom H, Martin A, McIntosh C, McMillan WO, Palumbi S. 1989. The simple fool’s guide to PCR. Honolulu: Department of Zoology, University of Hawaii. Kolbe JJ, Glor RE, Rodríguez Schettino L, Lara AC, Larson A, Losos JB. 2004. Genetic variation increases during biological invasion by a Cuban lizard. Nature 431:177–181 DOI 10.1038/nature02807. Kolbe JJ, Larson A, Losos JB, De Queiroz K. 2008. Admixture determines genetic diversity and population differentiation in the biological invasion of a lizard species. Biology Letters 4:434–437 DOI 10.1098/rsbl.2008.0205. Lafferty KD, Page CJ. 1997. Predation on the endangered tide-water goby, Eucyclogobius newberryi, by the introduced African clawed frog, Xenopus laevis, with notes on the frog’s parasites. Copeia 3:589–592 DOI 10.2307/1447564. Lavergne S, Molofsky J. 2007. Increased genetic variation and evolutionary potential drive the success of an invasive grass. Proceedings of the National Academy of Sciences of the United States of America 104:3883–3888 DOI 10.1073/pnas.0607324104. Lee CE. 2002. Evolutionary genetics of invasive species. Trends in Ecology & Evolution 17:386–391 DOI 10.1016/S0169-5347(02)02554-5. Lee CE, Gelembiuk GW. 2008. Evolutionary origins of invasive populations. Evolutionary Applications 1:427–448 DOI 10.1111/j.1752-4571.2008.00039.x. Librado P, Rozas J. 2009. DnaSP v5: a software for comprehensive analysis of DNA poly- morphism data. Bioinformatics 25:1451–1452 DOI 10.1093/bioinformatics/btp187. Lillo F, Dufresnes C, Faraone FP, Lo Valvo M, Stöck M. 2013. Identification and potential origin of invasive clawed frogs Xenopus (Anura: Pipidae) in Sicily based on mitochondrial and nuclear DNA. Italian Journal of Zoology 80:566–573 DOI 10.1080/11250003.2013.847502. Lillo F, Faraone FP, Lo Valvo M. 2010. Can the introduction of Xenopus laevis affect native amphibian populations? Reduction of reproductive occurrence in presence of the invasive species. Biological Invasions 13:1533–1541 DOI 10.1007/s10530-010-9911-8. Lillo F, Marrone F, Sicilia A, Castelli G, Zava B. 2005. An invasive population of Xenopus laevis (Daudin, 1802) in Italy. Herpetozoa 18:63–64. Lips KR, Brem F, Brenes R, Reeve JD, Alford RA, Voyles J, Carey C, Livo L, Pessier AP, Collins JP. 2006. Emerging infectious disease and the loss of biodiversity in a Neotropical amphibian community. Proceedings of the National Academy of Sciences of the United States of America 103:3165–3170 DOI 10.1073/pnas.0506889103. Lobos G, Measey GJ. 2002. Invasive populations of Xenopus laevis (Daudin) in Chile. Herpetological Journal 12:163–168. Lobos G, Mendez MA, Cattan P, Jaksic F. 2014. Low genetic diversity of the successful invasive African clawed frog Xenopus laevis (Pipidae) in Chile. Studies on Neotropical Fauna and Environment 49:50–60 DOI 10.1080/01650521.2014.912865. Mardulyn P. 2012. Trees and/or networks to display intraspecific DNA sequence variation? Molecular Ecology 21:3385–3390 DOI 10.1111/j.1365-294X.2012.05622.x.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 18/21 Measey GJ. 2004. Species account: Xenopus laevis. In: Minter LR, Burger M, Harrison JA, Braack HH, Bishop PJ, eds. Atlas and Red Data book of the Frogs of South Africa, Lesotho and Swaziland. Washington, D.C.: Smithsonian Institution Press, 266–267. Measey GJ, Channing A. 2003. Phylogeography of the genus Xenopus in southern Africa. Amphibia-Reptilia 24:321–330 DOI 10.1163/156853803322440781. Measey GJ, Rödder D, Green SL, Kobayashi R, Lillo F, Lobos G, Rebelo R, Thirion JM. 2012. Ongoing invasions of the African clawed frog, Xenopus laevis: a global review. Biological Invasions 14:2255–2270 DOI 10.1007/s10530-012-0227-8. Measey GJ, Vimercati G, De Villiers FA, Mokhatla MM, Davies SJ, Edwards S, Altwegg R. 2015. Frog eat frog: exploring variables influencing anurophagy. PeerJ 3:e1204 DOI 10.7717/peerj.1204. Nixon KC. 1999. The Parsimony Ratchet, a New Method for Rapid Parsimony Analysis. Cladistics 15:407–414 DOI 10.1111/j.1096-0031.1999.tb00277.x. Novak S, Mack R. 1993. Genetic variation in Bromus tectorum (Poaceae): comparison between native and introduced populations. Heredity 71:167–176 DOI 10.1038/hdy.1993.121. Peakall R, Smouse PE. 2012. GenAlEx 6.5: genetic analysis in Excel. Population genetic software for teaching and research–an update. Bioinformatics 28:2537–2539 DOI 10.1093/bioinformatics/bts460. Pearman PB, D’Amen M, Graham CH, Thuiller W, Zimmermann NE. 2010. Within- taxon niche structure: niche conservatism, divergence and predicted effects of climate change. Ecography 33:990–1003 DOI 10.1111/j.1600-0587.2010.06443.x. Peralta-García A, Valdez-Villavicencio JH, Galina-Tessaro P. 2014. African clawed frog (Xenopus laevis) in Baja California: a confirmed population and possible ongoing invasion in Mexican watersheds. The Southwestern Naturalist 59:431–434 DOI 10.1894/NBF-12.1. Posada D. 2008. jModelTest: phylogenetic model averaging. Molecular Biology and Evolution 25:1253–1256 DOI 10.1093/molbev/msn083. Prentis PJ, Sigg DP, Raghu S, Dhileepan K, Pavasovic A, Lowe AJ. 2009. Understanding invasion history: genetic structure and diversity of two globally invasive plants and implications for their management. Diversity and Distributions 15:822–830 DOI 10.1111/j.1472-4642.2009.00592.x. R Development Core Team. 2014. R: a language and environment for statistical computing. Vienna: the R Foundation for Statistical Computing. Available at http://www.R- project.org/. Rambaut A. 2014. FigTree: Tree Figure Drawing Tool. Version 1.4.2. Available at http: //tree.bio.ed.ac.uk/software/figtree/. Rambaut A, Drummond AJ, Suchard M, Xie W. 2013. Tracer: MCMC Trace Analysis Tool. Version 1.6.0. Available at http://tree.bio.ed.ac.uk/software/tracer/. Rebelo R, Amaral P, Bernardes M, Oliveira J, Pinheiro P, Leitão D. 2010. Xenopus laevis (Daudin, 1802), a new exotic amphibian in Portugal. Biological Invasions 12:3383–3387 DOI 10.1007/s10530-010-9757-0.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 19/21 Rius M, Darling JA. 2014. How important is intraspecific genetic admixture to the success of colonising populations? Trends in Ecology & Evolution 29:233–242 DOI 10.1016/j.tree.2014.02.003. Ronquist F, Huelsenbeck JP. 2003. MrBayes 3: Bayesian phylogenetic inference under mixed models. Bioinformatics 19:1572–1574 DOI 10.1093/bioinformatics/btg180. Rose W, Hewitt J. 1927. Description of a new species of Xenopus from the Cape peninsula. Transactions of the Royal Society of South Africa 14:343–346. Sakai AK, Allendorf FW, Holt JS, Lodge DM, Molofsky J, With KA, Baughman S, Cabin RJ, Cohen JE, Ellstrand NC, Mccauley DE, O’Neil P, Parker IM, Thompson JN, Weller SG. 2001. The population biology of invasive species. Annual Review of Ecology and Systematics 32:305–332 DOI 10.1146/annurev.ecolsys.32.081501.114037. Schliep KP. 2011. phangorn: phylogenetic analysis in R. Bioinformatics 27:592–593 DOI 10.1093/bioinformatics/btq706. Skerratt LF, Berger L, Speare R, Cashins S, McDonald KR, Phillott AD, Hines HB, Kenyon N. 2007. Spread of chytridiomycosis has caused the rapid global decline and extinction of frogs. EcoHealth 4:125–134 DOI 10.1007/s10393-007-0093-5. Stephens M, Donnelly P. 2003. A comparison of bayesian methods for haplotype reconstruction from population genotype data. American Journal of Human Genetics 73:1162–1169 DOI 10.1086/379378. Stephens M, Smith NJ, Donnelly P. 2001. A new statistical method for haplotype reconstruction from population data. American journal of human genetics 68:978–89 DOI 10.1086/319501. Stöver BC, Müller KF. 2010. TreeGraph 2: combining and visualizing evidence from different phylogenetic analyses. BMC Bioinformatics 11:7 DOI 10.1186/1471-2105-11-7. Tamura K, Stecher G, Peterson D, Filipski A, Kumar S. 2013. MEGA6: Molecular Evolu- tionary Genetics Analysis version 6.0. Molecular Biology and Evolution 30:2725–2729 DOI 10.1093/molbev/mst197. Thompson JD, Higgins DG, Gibson TJ. 1994. CLUSTAL W: improving the sensitivity of progressive multiple sequence alignment through sequence weighting, position- specific gap penalties and weight matrix choice. Nucleic Acids Research 22:4673–4680 DOI 10.1093/nar/22.22.4673. Tinsley RC, Loumont C, Kobel HR. 1996. Geographical distribution and ecology. In: Tinsley RC, Kobel HR, eds. The biology of Xenopus. Oxford: Oxford University Press, 35–59. Tinsley RC, McCoid MJ. 1996. Feral populations of Xenopus outside Africa. In: Tinsley RC, Kobel HR, eds. The biology of Xenopus. Oxford: Oxford University Press, 81–94. Voyles J, Young S, Berger L, Campbell C, Voyles WF, Dinudom A, Cook D, Webb R, Alford RA, Skerratt LF, Speare R. 2009. Pathogenesis of chytridiomycosis, a cause of catastrophic amphibian declines. Science 326:582–585 DOI 10.1126/science.1176765. Weldon C, De Villiers AL, Du Preez LH. 2007. Quantification of the trade in Xenopus laevis from South Africa, with implications for biodiversity conservation. African Journal of Herpetology 56:77–83 DOI 10.1080/21564574.2007.9635553.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 20/21 Weldon C, Du Preez LH, Hyatt AD, Muller R, Spears R. 2004. Origin of the amphibian chytrid fungus. Emerging Infectious Diseases 10:2100–2105 DOI 10.3201/eid1012.030804.

De Busschere et al. (2016), PeerJ, DOI 10.7717/peerj.1659 21/21

APPENDIX J

Measurements performed on each individual involved in morphometric comparisons.

M10

M12 M11

M2 M9

M3 M8

M7 M5 M4 M6

Measurements performed on each individuals involved in morphometric comparisons. M1 = Snout-vent length; M2 = Femur; M3 = Tibia; M4 = Tarses; M5 = 5th toe; M6 = Humerus; M7 = Radius; M8 = Carpes; M9 = 2nd finger; M10 = Length; M11 = Width; M12 = Heigth

Abstract

Because of the current global biodiversity decline, understanding the consequences of each threat on biodiversity is crucial for conservation biology. Invasive species are among the main threats at the global scale, and can locally imply harmful damages on ecosystems. Studying the phenomena driving the effects and potential for expansion of these species appears as a crucial element to assess their long terms impacts. In this study, we focused our efforts on an invasive population of the

African clawed frog, Xenopus laevis, in France, to bring insight about the interactions of this population with its environment and to study the changes in resource allocation to the life history traits, related to reproduction, survival and dispersal probabilities, during the range expansion of the population. We studied the diet in the French invasive population and in other invasive and native populations, and found that this species can expand by predating a narrow, as well as a broad, range of prey categories. We also detected an impact of X. laevis on the native amphibian community in

France. In the second section of the thesis, we reported a decrease in reproductive investment, and an increased dispersal allocation of resources at the range edge. We finally studied population dynamics and detected a lower survival probability and density at the range core. All these results combined suggest that the potential for long term impacts is important in France for X. laevis as well as in other areas where the species has been, or will be, introduced.