vol. 190, no. 5
the american naturalist
november 2017
Bioinvasion Triggers Rapid Evolution of Life Histories in Freshwater Snails Elodie Chapuis,1,2,3,* Thomas Lamy,2,4 Jean-Pierre Pointier,5 Nicolas Juillet,3 Adeline Ségard,2 Philippe Jarne,2 and Patrice David2 1. Institut de Recherche pour le Développement, Cirad, Université de Montpellier, Interactions Plantes-Microorganismes-Environnement, 34394 Montpellier cedex 05, France; 2. Centre d’Ecologie Fonctionnelle et d’Evolution, Unité Mixte de Recherche (UMR) 5175, CNRS– Université de Montpellier–Université Paul Valéry Montpellier–École pratique des hautes études (EPHE), 1919 route de Mende, 34293 Montpellier cedex 5, France; 3. UMR Peuplements végétaux et bioagresseurs en milieu tropical, Université de la Réunion–Centre de coopération internationale en recherche agronomique pour le développement, 7 Chemin de l’IRAT, Ligne Paradis, 97410 Saint Pierre, La Réunion, France; 4. Département de Sciences Biologiques, Université de Montréal, C.P. 6128, Succursale Centre-ville, Montréal, Quebec H3C 3J7, Canada; 5. Unité de Service et de Recherche 3278 CNRS–EPHE CRIOBE, Université de Perpignan, 66880 Perpignan cedex, France Submitted February 7, 2017; Accepted June 2, 2017; Electronically published September 5, 2017 Online enhancements: appendix with two parts. abstract: Biological invasions offer interesting situations for observing how novel interactions between closely related, formerly allopatric species may trigger phenotypic evolution in situ. Assuming that successful invaders are usually filtered to be competitively dominant, invasive and native species may follow different trajectories. Natives may evolve traits that minimize the negative impact of competition, while trait shifts in invasives should mostly reflect expansion dynamics, through selection for colonization ability and transiently enhanced mutation load at the colonization front. These ideas were tested through a large-scale common-garden experiment measuring life-history traits in two closely related snail species, one invasive and one native, cooccurring in a network of freshwater ponds in Guadeloupe. We looked for evidence of recent evolution by comparing uninvaded or recently invaded sites with long-invaded ones. The native species adopted a life history favoring rapid population growth (i.e., increased fecundity, earlier reproduction, and increased juvenile survival) that may increase its prospects of coexistence with the more competitive invader. We discuss why these effects are more likely to result from genetic change than from maternal effects. The invader exhibited slightly decreased overall performances in recently colonized sites, consistent with a moderate expansion load resulting from local founder effects. Our study highlights a rare example of rapid life-history evolution following invasion. Keywords: bioinvasions, phenotypic traits, competition, metacommunity, mollusc, Aplexa, Physa.
* Corresponding author; e-mail:
[email protected]. ORCIDs: Chapuis, http://orcid.org/0000-0003-2299-0992; Lamy, http://orcid .org/0000-0002-7881-0578; Juillet, http://orcid.org/0000-0002-0705-4541. Am. Nat. 2017. Vol. 190, pp. 694–706. q 2017 by The University of Chicago. 0003-0147/2017/19005-57540$15.00. All rights reserved. DOI: 10.1086/693854
Introduction Biological invasions have been a central topic for decades in ecology (Davis 2009) and, more recently, attracted the attention of evolutionary biologists (Facon et al. 2006; Sax et al. 2007). They represent natural experiments for understanding eco-evolutionary dynamics over short time periods (Lankau and Strauss 2011; Hanski 2012). The establishment and spread of invasive species pose an evolutionary challenge to the invasive species themselves, which enter novel environments and communities, but also to native species, which develop new competitive, mutualistic, or trophic interactions with the invaders (Sax et al. 2005; Davis 2009). Invaded habitats are therefore expected to be evolutionary hot spots. Particularly interesting situations occur when invasive species compete with related natives with similar lifestyles and traits, providing the opportunity to compare phenotypic trajectories associated with the invasive/native status (Greenlees et al. 2010; Stuart et al. 2014). In such situations, we envision two major sets of selective pressures. The first is due to exploitation or interference competition and should essentially affect native species. It should be prevalent when native and invasive species occupy similar niches and are at risk of competitive exclusion (Davis 2009; Li et al. 2015). Recent theory has highlighted that not only niche differences (Sax et al. 2005) but also lifehistory differences—for example, competition-colonization trade-offs in metacommunities (Chase and Leibold 2003; Calcagno et al. 2006; Li et al. 2015)—can promote species coexistence. Thus, interactions between invasive and native species may select for divergence in niche or life-history traits, that is, for a character shift (Leibold et al. 2004; Calcagno et al. 2006; Li et al. 2015) that may occur in the course of a
Rapid Evolution following Bioinvasion few generations (review in Whitney and Gabler 2008). Such changes are more likely to be observed in natives than in invaders. Poor competitors are indeed filtered out in the early phases of invasion, as they are not likely to successfully spread in the presence of natives (Sax et al. 2005; Davis 2009). Therefore, competition is probably asymmetric in most invasive-native pairs. In agreement with this idea, models predict that, although native competitors may modify the selection regime on invaders compared to what they experienced in their native area, trait shifts in introduced species will rarely be fast enough to allow a species to invade if its initial trait means does not allow a positive growth rate in the presence of native species. Even when invasive populations have lots of genetic variation, invasion will depend on trait evolution only if both native and invasive species happen to reside within a very narrow range of trait space (Jones and Gomulkiewicz 2012). This has been supported by long-term surveys of plant communities (Li et al. 2015). In contrast, native species are progressively exposed to an increasing abundance of invaders, giving them more time to evolve, if they are not wiped out. Examples of rapid evolution in natives (review in Strauss et al. 2006; Sax et al. 2007; Westley 2011) indeed highlight changes in traits that allow them to escape negative interactions (generally, predation) with invaders. For example, the cane toad invasion in Australia selected for delayed breeding and smaller metamorphs in native frogs, decreasing their probability of eating the poisonous toad tadpoles (Greenlees et al. 2010; Shine 2012). The second major set of selective pressures concerns the invaders and is related to population expansion. First, traits favoring successful colonization, such as high dispersal and fast reproduction, should transiently occur at a higher frequency in expanding populations at the invasion front before being replaced by types adapted to stable established populations when the invasion front progresses (Phillips 2009; Kubisch et al. 2013; Hargreaves and Eckert 2014). The few available empirical examples seem to confirm this prediction. For example, introduced cane toads in Australia have longer legs and higher dispersal ability at the invasion front than in long-established populations (Phillips et al. 2006; Phillips 2009). Similarly, recently invaded populations of Spartina alterniflora show earlier reproduction, increased reproductive effort, and higher self-compatibility than established ones (Davis 2005). Second, founder effects may affect local populations at the colonization front (Baker and Stebbins 1965), even if genetic diversity in the whole invaded area may be as high as in the origin area (Roman and Darling 2007; Facon et al. 2008). Local bottlenecks transiently enhance genetic drift at the invasion front (Excoffier and Ray 2008), not only affecting neutral genetic diversity but also allowing some deleterious alleles to rise in frequency (Kirkpatrick and Jarne 2000) and to decrease the associated fitness traits, the so-called expansion load (Peischl and Ex-
695
coffier 2015). This should result in a nonadaptive change in traits, leading to fitness loss in recently occupied sites when compared to more anciently colonized ones. Examples of this process have been reported in human populations (Peischl et al. 2013) and in plants (Ellstrand and Schierenbeck 2000). In summary, (i) selective pressures should generate contrasts in traits between sites that have not been invaded or have recently been invaded on one side and sites with a longer history of invasion on the other side. In native species, selection should predominantly induce a character shift in response to the presence of invasive species, while it should (transiently) favor colonization ability in invasives; (ii) in invasives, trait values associated with depressed fitness are expected at the colonization front in comparison to settled populations. It should be highlighted that most studies on character shift in native and invasive species have focused on trophic (e.g., predation) or host-parasite interactions (Torchin et al. 2003; Colautti et al. 2004), while fewer have been interested in competition (Strauss et al. 2006; Wilson 2014). In addition, they were generally not concerned with closely related invasive-resident species pairs. An exception is a recent study showing that the competition exerted by an invasive lizard resulted in adaptive change in limb morphology and subsequent niche shift in a congeneric resident species in less than 20 generations (Stuart et al. 2014). Importantly, while life-history differences and not only niche differences have been identified as important drivers of species coexistence in communities, we still lack clear examples of life-history trait changes resulting from competitive interactions between closely related native and invasive species. We test our predictions using two closely related freshwater snail species that co-occur in the Guadeloupe Archipelago (Lesser Antilles). The native is Aplexa marmorata, and the invader is Physa acuta. They feed on the same resource, but A. marmorata reproduces faster and has a shorter life span than P. acuta, suggesting that their coexistence is more likely to depend on life-history than on trophic niche differences. Moreover, P. acuta excludes A. marmorata in laboratory cocultures (E. Chapuis, T. Lamy, J. P. Pointier, N. Juillet, A. Ségard, P. Jarne, and P. David, unpublished data). On the whole, A. marmorata exhibits traits of efficient colonizers, while P. acuta seems more efficient at resource exploitation and competition. Based on a long-term survey (Lamy et al. 2013a), we know the arrival date of P. acuta in a large number of sites, allowing us to contrast (i) populations of A. marmorata that have been invaded by P. acuta for various durations and (ii) populations of P. acuta at the colonization front versus those settled for several years. We studied several life-history traits in such populations (16 in P. acuta and 14 in A. marmorata) under common-garden conditions.
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The American Naturalist Methods Sites and Species Studied
Our study system includes 257 freshwater sites (mostly ponds) in the Guadeloupe Archipelago (fig. 1), located on Grande-Terre Island (231 sites) and the nearby island of Marie-Galante (26 sites). These sites have been surveyed yearly (January–February) from 2001 to 2010 to assess snail assemblages—29 species that together account for the largest part of the macrobenthos in these ecosystems (Lamy et al. 2012, 2013a). Only half are native as a result of repeated introductions over the past 50 years (Pointier 2008). The sampling methodology is presented in detail in Lamy et al. (2012, 2013a, 2013b). The variables recorded per site in these surveys that are of interest in this study are the pres-
ence/absence and density of each snail species (density was recorded as classes on a semiquantitative log scale, that is, 0 p no individuals; 1 p !1 individual per m2, 2 p 1–5 individuals per m2, 3 p 5–10 individuals per m2, . . .) and several ecological variables, including site size, connectivity, hydrological stability, and percent cover by aquatic vegetation and by the 21 most prevalent plant species. We here focus on the two Physidae species of this snail community, the native Aplexa marmorata and the introduced Physa acuta. The two species are closely related (Wethington and Lydeard 2007) and exhibit similar ecology and shell phenotype (e.g., adult shell length of up to 15 mm; Pointier 2008) but do not hybridize. Aplexa marmorata is a Neotropical species, while P. acuta originates from the eastern-central part of the United States (Bousset et al.
SJ BF
BE MA
LU SM LP "
DA JC TER
"
AL
PC CE
GO
PB FBE FT "
"
10 km
GB EB
Figure 1: Distribution of the sites sampled in Guadeloupe (top, Grande-Terre; bottom, Marie-Galante) in Physa acuta (circles) and Aplexa marmorata (squares). N (never or newly invaded) populations of Aplexa marmorata are indicated with a box. See table 1 for site acronyms.
Rapid Evolution following Bioinvasion 2014; Lydeard et al. 2016). However, P. acuta has markedly increased its range over the past two centuries and is now cosmopolitan. The two species currently occur sympatrically in the Lesser Antilles, South Africa, and Ivory Coast (Bony et al. 2008). In Guadeloupe, A. marmorata is the most prevalent freshwater snail, having been found at least once in 95.8% of the surveyed sites. Physa acuta was first detected in two sites of Grande-Terre in 2001 (it was reported in previous surveys since 1972 but only in the city of Pointe-àPitre, Grande-Terre) and quickly spread to currently occupy ∼60% of freshwater habitats of Grande-Terre. The two species have been coexisting at the same sites for 0–10 years (as of 2010). These two hermaphroditic species exhibit contrasted mating systems. Aplexa marmorata self-fertilizes at rates exceeding 90% in natural populations (Dubois et al. 2008), while the selfing rate is generally not different from zero in populations of P. acuta (Henry et al. 2005). In both species, egg capsules include from a few to a few tens of eggs inserted in a gelatinous matrix (Jarne et al. 2010) and are laid on solid substrate (e.g., box walls in the laboratory). Juveniles hatch after about a week. Sexual maturity is reached in about 4– 5 weeks, and individuals can live up to 6 months under our laboratory conditions. Growth is continuous and adult shell length ranges from 10 to 15 mm. About six generations are completed per year under natural conditions. Common-Garden Experiment We conducted a common-garden experiment to measure life-history traits in first- or second-generation laboratory offspring of wild-caught individuals. Note that this allows us to interpret results in terms of transgenerational effects (maternal or genetic) and to exclude plastic responses to different habitats. Nineteen sites were sampled on Guadeloupe and Marie-Galante in January 2010, in which one or both species were collected (fig. 1; table 1). Physa acuta was sampled at 16 sites and A. marmorata at 14 sites. The two species co-occurred in 11 of these 19 sites (table 1). Based on knowledge of site occupancy from 2001 to 2009, A. marmorata populations were sampled in order to contrast two site categories: not invaded by P. acuta (naive populations) versus invaded for more than 2 years (experienced populations; mean number of years since invasion p 4:77 5 1:5 SD). However, a few sites of the first category were reached by P. acuta between 2009 and 2010, so that this category became “not or recently invaded.” The recently invaded populations of A. marmorata are unlikely to show a detectable evolutionary response to the presence of P. acuta because of the limited number of generations. Colonization indeed occurs mainly during the rainy season owing to temporary water connections among sites, that is, less than 6 months (three generations) before sampling, during which
697
densities of P. acuta were likely still low. Thus, we kept the analysis as planned (contrasting two groups) after having checked that there were no significant differences between uninvaded and recently invaded populations within the naive category and no significant effect of time since invasion (as a continuous variable) within the experienced category (see appendix, pt. A and table A1; appendix and tables A1– A3 available online). In P. acuta, we similarly defined two categories of populations: populations from sites invaded for less than a year (colonization front) versus sites behind the colonization front invaded for at least 3 years (core populations). Models considering time since invasion as a continuous linear factor did not explain phenotypic variance as well as models using the categorical front/core factor (table A1). Therefore, we report the results of the latter models. About 20 mature individuals (generation 0, G0) were randomly collected per species and site by three people exploring the entire sites. They were brought back to the laboratory in Montpellier and kept isolated in plastic boxes (75 ml). Egg capsules from 10–15 G0 individuals per population were collected over a week. On the whole, the experiment was based on 700 G1 individuals in P. acuta and 550 in A. marmorata. The offspring of a given G0 mother will be referred to as a family. Each family consisted of 3–5 G1 juveniles. We recorded reproductive traits of the G1 individuals of both species under their preferred mating system. Aplexa marmorata is a predominantly selfing hermaphrodite, in which reproduction in isolation is associated with no significant inbreeding depression and no reproductive delay (Escobar et al. 2011). Individuals were therefore isolated 15 days after hatching (i.e., before sexual maturity) until the end of the experiment, leaving self-fertilization as the only possible reproductive mode. In contrast, P. acuta is a preferential outcrosser, in which enforcing self-fertilization delays reproduction and reduces offspring survival (Escobar et al. 2011). The G1 individuals were also isolated 15 days after hatching and remained isolated, except for three 2-hour periods per week during which they were grouped (10 individuals, marked with paint dots). This protocol ensures that individuals reproduce through outcrossing while limiting density effects (Tsitrone et al. 2003; Escobar et al. 2011). Individuals were monitored for reproduction and survival over 6 months (more details below), a point at which mortality began to markedly increase. Snails were kept under a 12L∶12D photoperiod at a room temperature of 247C 5 17C and fed ad lib. with boiled ground lettuce. Twice a week, water was changed and the position of rearing boxes was randomly moved. Phenotypic Traits The G1 individuals were monitored throughout their life cycle (up to 6 months) in order to measure several traits
BE MA DA EB GB PB LP BF FT SJ LU JC PC SM AL CE FBE GO TER
Acronym Never Never 2010 2010 2010 2010 2008 2007 2007 2007 2006 2005 2005 2005 2004 2004 2004 2004 2004
261729 21.06 261728010.0800 261728053.3400 261715033.3600 261714027.4800 261727056.2200 261726026.7600 261729017.9900 261724056.2200 261727053.100 261724055.7400 261731059.700 2617290 50.5800 2617240 17.0900 2617170 23.2200 2617290 27.0600 2617260 10.5600 2617280 37.1400 2617300 11.0900
16725 15.06 16723018.8400 16717021.6600 1575706.4800 15757039.3000 16713038.7600 16720027.4200 16727054.3000 1671303.6600 16729016.8600 16722048.3600 16716013.9200 16715040.3800 16721030.3600 1671907.0800 16714055.0200 16713035.1000 16713026.2200 16715014.6400
00
First detection
0
Longitude
0
00
Latitude
E E E
E E E
E
N N E E
N N N
A. marmorata status
C C C C C C C C C C C C
F F F F
P. acuta status
x x x
x x x
x
x x x x
x x x
A. marmorata
x x x x x x x x x x x x
x x x x
P. acuta
Note: The year during which P. acuta was first detected is given in the “first detection” column (“never” indicates that P. acuta was never detected). The corresponding invasion status is given for each species (N for naive, E for experienced, F for front, and C for core). Crosses in the last two columns indicate which species were sampled.
Belin Malakoff David Etang Bambou Ravine Grand-Bassin Port Blanc Laroche Picard Beaufond Fond Thezan Saint Jacques Lubeth Jacquot Prairie Pont de Caraque Sainte Marguerite Nord Alleaume Nord Caraque Est Fond Bertrand Goudenave Terrasson
Site
Table 1: Nineteen sites in which Aplexa marmorata (native) and Physa acuta (invasive) were sampled and their GPS coordinates (using World Geodetic System 84)
Rapid Evolution following Bioinvasion (table 2), as in previous studies (see, e.g., Chapuis et al. 2007; Escobar et al. 2011). Shell length and width were measured 8 and 32 days after individuals were isolated. From these values, we also estimated a shape parameter, namely, the length-over-width ratio. Shell length was also recorded at sexual maturity, estimated as the date of first egg laying and varying among individuals from 25 to 35 days after hatching. Measurements were performed to the nearest 0.01 mm using a binocular microscope. We also recorded the age at both sexual maturity (first egg laying) and death. Individual fecundity was estimated as the number of eggs and of egg capsules laid over 15 days after sexual maturity. We also estimated the area of eggs and of egg capsules (squashed between two glass plates to a constant thickness) as a measure of reproductive investment. This was performed by photographing three capsules per individual and three eggs per capsule, followed by image analysis using ImageJ 1.46r (Scheinder et al. 2012). Another 4–5 capsules were kept per G1 parent during the third week after sexual maturity, and the survival of G2 juveniles was estimated 15 days after egg laying as the ratio of surviving juveniles over the total number of eggs. Statistical Analyses of Phenotypic Data Linear mixed models and generalized linear mixed models were used to analyze the 11 life-history traits considered (see table 2) separately in the two species using lme4 (Bates et al. 2014) in R (R Core Team 2015). Fecundity was log
699
transformed, while age at sexual maturity and at death were square-root transformed to follow a normal distribution. Survival data were analyzed assuming a binomial distribution. We tested for the effect of population status as a fixed effect and of both the population factor, nested within status, and the family factor, nested within population, as random effects. Factors were tested using log-likelihood ratio tests (LRTs) and model simplification. For the linear mixed models, we conducted LRTs for the fixed effect using a maximum likelihood (ML) model, while random effects were tested using a restricted maximum likelihood (REML) model, as advised in Zuur et al. (2009). Multivariate analyses were performed to summarize the information conveyed by these 11 traits to a reduced number of dimensions and to analyze trait contribution to the a priori classification of populations according to their status (naive/experienced in A. marmorata and front/core in P. acuta). A principal component analysis (PCA) was carried out with the package ade4 (Dray and Dufour 2007) in R for each species separately: fecundity was log transformed, juvenile survival was arc sine transformed, and age at sexual maturity and at death were square root transformed. A linear discriminant analysis further allowed evaluating the contribution of individual traits to status difference in both species. Checking for Undesired Effects Several effects may interfere with the interpretation of phenotypic change in the two species studied. We first checked
Table 2: Mean values of the 11 life-history traits studied in individuals of Aplexa marmorata from naive (N) and experienced (E) populations (definitions in text) Trait Age at death (days) Age at sexual maturity (days) Capsule area (mm2) Egg area (mm2) N eggs N egg capsules Length (at 8 days; mm) Length (at sexual maturity; mm) Length (at 32 days; mm) Length/width (at 32 days) Juvenile survival (%)
Acronym AGE.DEATH AGE.MAT AREA.CAP AREA.EGG N.EGG N.CAP
Naive 57.26 (2.692) 29.19 (.812) .346 (.077) .0067 (.00045) 117.60 (3.934) 12.13 (1.155)
Experienced
Status
48.97 (4.196)
.135
27.22 (.663) .206 (.011) .0054 (.00012) 148.53 (6.576) 15.64 (.317)
Population
Family
PCA1
PCA2
.012
!.001
.30
2.25
.082 .037 .014 .013 .015
!.001 .641 .001 .232 .147
!.001 !.001 !.001 !.001 !.001
.50 .87 .91 2.24 2.76
.55 2.12 .02 2.81 2.41
L8
4.03 (.164)
4.18 (.183)
.561
!.001
!.001
.09
2.77
L.MAT
6.89 (.265)
6.51 (.0811)
.079
!.001
!.001
.66
2.40
L32
8.04 (.371)
7.73 (.196)
.380
.453
!.001
.69
2.56
2.04 (.027) 64.52 (5.555)
2.02 (.009) 85.44 (1.796)
.255 !.001
.425 !.001
!.001 !.001
.58 2.70
2.05 2.18
R32 JUV.SURV
Note: Standard errors calculated using a mixed model to take into account population and family effects are given in parentheses. The other columns report the significance associated with the status, populations (nested within status), and family (nested within population) effects from the likelihood ratio tests in the models. Boldface indicates P values lower than .05. The trait coordinates on the first two axes of the principal component analyses (PCA1 and PCA2) are also given. More details on statistical analyses are in the text.
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The American Naturalist
that sites that have been uninvaded or recently invaded (!1 year) by P. acuta and sites that have a longer history of invasion (12 years) had similar ecological characteristics in terms of vegetation cover, hydrological stability, size, and connectivity using a multivariate analysis of variance. Moreover, we tested in both species whether populations belonging to the same category were not geographically closer to each other than populations belonging to different categories using a Mantel test. Testing for Environmental and Demographic Consequences of P. acuta Invasion We also evaluated whether the invasion of P. acuta may have been correlated to environmental changes or to the demography of non-Physid species. Such changes sometimes occur, as exemplified by the introduction of the snail Marisa cornuarietis in some Guadeloupe ponds 25 years ago, resulting in a sharp decrease of pond cover by the water lily Nymphaea ampla and in a collapse of the metapopulation of the snail Biomphalaria glabrata (Pointier and David 2004). We tested whether the composition of aquatic vegetation at the 19 sites has significantly changed after the invasion of P. acuta. We created a binary variable accounting for the invasion of P. acuta between 2001 and 2010 at each site (1 in invaded sites and 0 otherwise) and tested for its effect on Hellinger-transformed vegetation composition by redundancy analyses (RDA) while controlling for any natural compositional differences among sites and any linear
temporal trend. This analysis was conducted with the library vegan in R. In order to evaluate whether the demography of non-Physid snail species was affected by the occurrence of P. acuta, we used demographic models allowing for estimation of site occupancy that were specifically developed for the Guadeloupe metacommunity (Lamy et al. 2013a). We here estimated site occupancy of the eight most common snail species besides Physids over the 2000–2010 period. A further important aspect, especially for evaluating the origin of a possible phenotypic shift in the native species, is whether the invasive species actually affects its demography in the Guadeloupe metacommunity. This was done using data from our long-term survey, more specifically using the density records (see above). In order to test whether the invasion of P. acuta has an influence on the demography of A. marmorata, we estimated the temporal variation of the density of the latter species as a function of the number of years since the first occurrence of P. acuta. This was done using a linear model with site and year as random factors, both in all 257 sites and in the subset of 19 sites sampled for the common-garden experiment. Results The mean trait values per population are reported in table 2 for Aplexa marmorata and in table 3 for Physa acuta. In A. marmorata, eight traits out of 11 exhibited significant differences (P ! :05), or near significant differences
Table 3: Mean values of the 11 life-history traits studied in individuals of Physa acuta from front (F) and core (C) populations (definitions in text) Trait Age at death (days) Age at sexual maturity (days) Capsule area (mm2) Egg area (mm2) N eggs N egg capsules Length (at 8 days; mm) Length (at sexual maturity; mm) Length (at 32 days; mm) Length/width (at 32 days) Juvenile survival (%)
Acronym AGE.DEATH AGE.MAT AREA.CAP AREA.EGG N.EGG N.CAP
Front 74.45 (4.637) 31.49 (.545) .274 (.038) .0045 (.00014) 136.08 (5.603) 8.13 (.681)
Core 79.97 (1.603) 30.56 (.167) .319 (.027) .0053 (.00039) 182.13 (6.952) 11.76 (.565)
Status
Population
Family
PCA1
PCA2
.054
.019
!.001
2.01
.33
.034 .329 .242 .012 .015
.231 .001 !.001 .999 .858
!.001 !.001 !.001 !.001 !.001
.66 2.15 2.06 2.89 2.71
.46 .73 .69 2.04 2.18
L8
3.84 (.101)
3.96 (.056)
.178
.012
.003
2.73
2.20
L.MAT
6.09 (.104)
6.24 (.089)
.341
.102
!.001
2.47
.48
L32
7.13 (.082)
7.21 (.112)
.660
.213
!.001
2.64
.36
1.75 (.006) 55.39 (5.000)
1.75 (.006) 52.22 (3.360)
.792 .382
.596 .324
1 !.001
2.09 2.05
.34 2.30
R32 JUV.SURV
Note: Standard errors calculated using a mixed model to take into account population and family effects are given in parentheses. The other columns report the significance associated with the status, population (nested within status), and family (nested within population) effects from the likelihood ratio tests in the models. Boldface indicates P values lower than .05. The trait coordinates on the first two axes of the principal component analyses (PCA1 and PCA2) are also given. More details on statistical analyses in text.
Rapid Evolution following Bioinvasion
4
A
AGE.MAT
Axis 2 (21.1%)
2
AREA.EGG
0
R32
Population N E
AGE.DEATH
-2
-4
-6
-4
-2
0
2
4
Axis 1 (39.3%)
B 6 AREA.CAP
Axis 2 (18%)
4
AREA.EGG
L.MAT
2
0
AGE.MAT
L32
R32
AGE.DEATH
Population F C
N.EGG N.CAP L8
-2
JUV.SURV
-3
0
3
Axis 1 (26.8%)
C AREA.CAP
Axis 2 (19.8%)
5
2.5
AREA.EGG
L.MAT
AGE.MAT
L32 R32
AGE.DEATH
0
PaF PaC
L8
N.EGG
-2.5
JUV.SURV N.CAP
-5
Population AmN AmE
-2.5
0
Axis 1 (30.6%)
2.5
5
701
(P ! :10), between sites newly or never colonized by P. acuta (naive populations) and sites invaded for more than 2 years (experienced). Aplexa marmorata individuals from experienced populations tended to start laying eggs earlier (27.2 vs. 29.2 days, P p :082) and laid more eggs (148.5 vs. 117.6, P p :013) and more egg capsules (15.6 vs. 12.1, P p :015) than individuals from naive populations. They also laid significantly smaller eggs and capsules, exhibited much higher juvenile survival (85.4% vs. 64.5%, P ! :001), and tended to die earlier. For size and shape, only the body length at sexual maturity differed between statuses, with a marginally significant smaller value in experienced than in naive populations (6.5 vs. 6.9, P p :079). In P. acuta, only three traits exhibited significant differences between front and core populations (table 3): individuals from core populations laid more eggs (182.1 vs. 136.1, P p :012) and more capsules (11.8 vs. 8.1, P p :015) and started to reproduce earlier (30.6 vs. 31.5 days, P p :034). No differences were found for body size and shape, and only slight tendencies toward longer life span were observed in core populations (table 3). We also observed significant variation among populations for seven traits in A. marmorata and for four traits in P. acuta and, more importantly, significant variation among families for all traits in both species, except on shell shape (i.e., length/width ratio) at 32 days in P. acuta (tables 2, 3). In A. marmorata, the first two axes of the PCA on the 11 traits explained more than 60% of the variance. On the first axis, juvenile survival and fecundity contributed negatively, while traits related to egg/capsule area and size parameters contributed positively (fig. 2A; table 2). The centroids of the naive and experienced populations were clearly separated along the first PCA axis for A. marmorata. In P. acuta, the first two axes explained only 45% of the variance (fig. 2B). As in A. marmorata, fecundity contributed negatively to the first axis (fig. 2B). Size parameters and egg/capsule area contributed negatively to this first axis (table 3). The centroid of the front and core categories were not clearly apart from each other. When data from the two species were considered together (fig. 2C), the same pattern appeared, with a clear separation between naive and experienced centroids (A. marmorata) and a limited one between front and core populations (P. acuta). The discriminant analysis between the naive and experienced populations in A. marFigure 2: First two axes of the principal component analyses conducted on the 11 variables (see tables 2, 3 for traits acronyms). Each dot is an individual. A, Aplexa marmorata (Am). B, Physa acuta (Pa). C, Aplexa marmorata and Physa acuta together. A, B, C, The contribution of each trait (variable) on the two axes is represented by an arrow. Individuals from the naive (N), experienced (E), front (F), and core (C) population groups are represented by different symbols. The centroids of each group are indicated as larger versions of the symbol referring to the relevant group.
Discussion Character Shift in the Native Species We predicted that native species should evolve to better tolerate the negative effects of competition by a related invader through character shift (Kolbe et al. 2004; Pfennig and Pfennig 2009). Life-history traits are likely to be involved in such shifts because of their potential role in promoting species coexistence. Indeed, modern metacommunity theory predicts that two species might coexist even if they exploit the same resource, by virtue of life-history trade-offs; for example, a good competitor/bad colonist can coexist with
2.0 1.8 1.6 1.4
morata was significant (permutation test, P ! :0001). Capsule size and fecundity parameters contributed to 32% and 29.5% of the variance on the discriminant axis. Age at sexual maturity and early survival contributed to a lower extent (15% and 7.5%). No significant difference was detected between the front and core sites in P. acuta. Uninvaded or recently invaded sites and sites with a longer history of invasion did not differ in terms of vegetation cover, hydrological stability, size, and connectivity, whether we considered A. marmorata populations only (Pillai’s trace p 0:124, F 1, 12 p 0:318, P p :860) or P. acuta ones (Pillai’s trace p 0:340, F 1, 14 p 1:413, P p :293). Moreover, populations belonging to the same invasion category were not geographically closer to one another than populations belonging to different categories in both A. marmorata (Mantel test; r p 0:02 and P p :346 or r p2 0:07 and P p :701 when the two sites from Marie-Galante were included or not, respectively) and P. acuta (Mantel test; r p 0:25 and P p :074 or r p2 0:16 and P p :942), suggesting that the sites chosen were randomly distributed over the islands with regard to invasion status. The invasion of P. acuta did not significantly affect the composition of aquatic vegetation (RDA R2adj p2 0:001, F 1, 160 p :847, P p :480), which nevertheless considerably varied among sites (R2adj p 0:286, F 18, 160 p 5:069, P ! :001). These results suggest that aquatic vegetation did not systematically vary after invasion by P. acuta. Moreover, the average site occupancy showed little variation over the 2000–2010 period in seven of the eight snail species considered (the most common ones) using our demographic approach. The only exception was Drepanotrema surinamense, with a slight increase until 2006 followed by a slight decrease. Overall, the snail metacommunity was not significantly affected by the spread of P. acuta. In sites invaded by P. acuta, the density of A. marmorata progressively decreased over years (fig. 3), with a highly significant effect both over all 257 sites (P ! :001) and over the 19 sites sampled for the common-garden experiment (P p :035).
Density score of A. marmorata
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0
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Years since first observation of P. acuta Figure 3: Density of Aplexa marmorata as a function of the time (in years) after site invasion by Physa acuta in all sites studied in the Guadeloupe metacommunity. Value 0 represents sites that have not been invaded, and all sites that have been invaded for more than 4 years were pooled because of a low number of sites in this case. The standard error is reported for each mean density.
a good colonist/poor competitor (Leibold et al. 2004; Calcagno et al. 2006). In parallel, the theory of life-history evolution predicts that any change in the overall demography and age structure can impose selection on how organisms allocate energy to different life-history traits (Charlesworth 1994; Roff 2002). It is often considered that invasive populations at the colonization front are growing in a context of relaxed competition, with selection favoring early and abundant reproduction at the expense of adult survival (Davis 2009). However, when the invader faces a related native species, competition may occur. If it results in a uniform increase or decrease in mortality at all ages, it does not change selection on life-history traits. However, if competition differentially affects one stage (e.g., juvenile survival or late fertility), then selection may favor more rapid (respectively, slower) life histories (Caswell 2007). Thus, it is difficult to predict which life history will be positively selected in a native species recently exposed to competition with an invasive congener without knowing the details of how competition occurs in nature. Our data suggest that P. acuta exerts a strong competition pressure on A. marmorata. Indeed, the average density of the latter quickly decreased in invaded sites after the arrival of P. acuta (fig. 3), while the community composition was otherwise not affected. Moreover, P. acuta excludes A. marmorata in cocultures under laboratory conditions (E. Chapuis, T. Lamy, J. P. Pointier, N. Juillet, A. Ségard, P. Jarne, and P. David, unpublished data). Although this is consistent
Rapid Evolution following Bioinvasion with direct competition between the two Physid species, apparent competition (Muller and Godfray 1997) cannot be excluded as a potential cause of this decline, calling for more detailed studies. For example, we do not know which life-history stage of A. marmorata is most affected by the interaction with P. acuta. Nevertheless, we detected rapid evolution in lifehistory traits of the native A. marmorata as early as 3–7 years after invasion, consistent with the general prediction that invasions may trigger life-history evolution in natives. In experienced populations, A. marmorata individuals show higher juvenile survival than in naive populations, begin to lay eggs slightly earlier, and produce more and smaller eggs packaged in many small capsules. Its life history, therefore, seems to have moved toward improved early performances (early production of more juveniles) at the expense of egg size (and, therefore, of hatching size, as development occurs without food intake within eggs) and, to a lower extent, of adult life span. It may be counterintuitive that the native species shifts toward a strategy that is often considered a hallmark of colonization-oriented and invasive species (Davis 2005). For freshwater snails occupying unstable habitats with cycles of drought/flood, such a strategy, in association with high selfing rates, might allow them to cope with environmental variation by efficiently recolonizing newly available habitats (Chapuis et al. 2007). However, the arrival date of P. acuta is not correlated to habitat instability, so the latter is not responsible for the differences between naive and experienced populations of A. marmorata. In addition, P. acuta does not show traits typical of a colonist strategy, compared to A. marmorata. On the contrary, it exhibits a slower life history with lower juvenile survival and later onset of reproduction, compensated by higher fecundity and longer life span. It is therefore likely that A. marmorata, by evolving toward a fast-reproducing phenotype with high juvenile survival, minimizes its interactions with the more competitive P. acuta, especially at the beginning of the dry season when snail communities are regrowing (Pointier 2008; Lamy et al. 2012). Aplexa marmorata may be able to quickly exploit relatively ephemeral patches of abundant resources within sites, before P. acuta becomes too abundant in these patches. Our work adds to the few previous studies supporting directional evolutionary change in the presence of an invasive competitor, for example, the change in limb length allowing Anolis lizards to exploit different perching heights in the vegetation as a result of invasion (Stuart et al. 2014). These studies are based on the same kind of evidence and share the same limits. While the observed character shifts are consistent with an adaptive hypothesis based on some coexistence theory (minimization of niche overlap through divergence in perching height in Anolis lizards; in our case, divergence along a competition-colonization axis through changes in life-history traits), the effectiveness of the as-
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sumed mechanisms of coexistence themselves is hard to test in the field. Accumulating more examples of character change in native species in response to invasion by related competitors would be a way to evaluate the importance of life-history traits and habitat or diet variation in species coexistence. An interesting question arising from our study is whether the observed change in A. marmorata life-history traits can be called an ecological character displacement (ECD; Schluter 2001; Pfennig and Pfennig 2009; Stuart and Losos 2013). Long-term surveys of bioinvasions, coupled with commongarden approaches, offer an ideal situation to evaluate Schluter’s six criteria for ECD (Stuart and Losos 2013). However, such displacements have generally been considered for morphological traits, for example, beak shape in Darwin finches (Stuart and Losos 2013; Lamichhaney et al. 2016) or skeletal traits in sticklebacks (Miller et al. 2015), which have more direct connection to resource use than life-history traits. Further studies are required to test Schluter’s fourth criterion (changes in phenotypes match ecological shifts in resource use), but our study could qualify as an example of ECD based on the remaining five criteria. In a wider perspective, generalizing ECD to life-history traits, and studying them in a larger number of cases, is certainly a worthwhile goal. Transitory Fitness Loss in the Invasive Species? As explained in the introduction, competition is expected to more strongly affect the native than the invasive species. In the invasive P. acuta, we nevertheless observed some differences between newly colonized populations at the invasion front and older, core populations. In core populations, the reproductive effort was slightly higher with an earlier initiation of reproduction and higher egg production, but life span and egg size were not different. These differences cannot be interpreted as a consequence of differences in coexistence time with the native species. Indeed, although front populations of P. acuta are recent, they have been founded by migrants originating from core populations, which have been exposed to competition with A. marmorata (present everywhere on the island; see “Methods”). These differences may have two evolutionary origins. First, reaching a new site may require particular traits (e.g., favoring dispersal and fast population growth), so that front populations concentrate individuals with this particular syndrome and, subsequently, evolve back to the optimum typical of stable (core) populations. This is the case, for example, in the cane toad, which shows increased leg length at the invasion front in Australia (Phillips et al. 2006; Phillips 2009). Our case does not fit this scenario because fast and intense reproduction would be more expected in front populations of P. acuta (in which fast population growth is required) than in core populations—while the reverse is observed.
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Second, new sites are initially founded by few individuals, leading to reduced genetic diversity and possible consequences on fitness traits (i.e., expansion load; Peischl and Excoffier 2015). It may take some time to recruit enough genetic variation through mutation or immigration to recover appropriate levels of genetic diversity (e.g., Facon et al. 2008). We might, therefore, expect that performances slightly improve as populations age, a hypothesis consistent with our observation of increased reproductive effort with no apparent cost on other traits. Note, however, that the magnitude of this effect remains modest, which is in agreement with previous work showing slightly lower performance of relatively isolated populations of P. acuta compared to more connected and genetically diverse ones in a European metapopulation (Escobar et al. 2008). Transgenerational Effects: Genetically Based or Maternal Environment? We found significant differences between population status for several life-history traits in both the native (naive vs. experienced) and the invasive (front vs. core) species, and these differences were much more pronounced in the native species. Given our design, they can derive from either maternal or genetic effects—collectively referred to as transgenerational effects. Previous work conducted on life-history traits in Physids, all in P. acuta, provide evidence for both possibilities. Abundant genetic variation was, for example, found by Noël et al. (2016, 2017). Weak maternal effects were detected by Escobar et al. (2008), and transgenerational plasticity in reaction to predator cue were found by Luquet and Tariel (2016). Maternal effects would here result from differential exposure to environmental conditions, but we did not find environmental parameters systematically differing between naive and experienced populations in A. marmorata and between front and core populations of P. acuta. Moreover, maternal effects should also be expressed more strongly in early expressed traits (see Räsänen and Kruuk 2007; Charmantier et al. 2014), that is, here early in the life cycle of individuals. It means also that they should diminish over time and thus be less strong in G2 individuals (grandmaternal effects) than in G1 individuals. We observed the contrary in A. marmorata: traits measured in G1 juveniles before reproduction (growth and shell shape traits) are not different between statuses, while significantly different traits were measured in mature G1 and in G2 (postreproductive traits, egg traits, and juvenile survival). The strongest status effect was indeed on survival of G2 juveniles. Moreover, we measured a few other traits in G2 offspring of some of the populations of both snail species, as part of a separate experiment (see appendix, pt. B). For the two traits in common with this experiment (juvenile shell length and number of egg capsules laid
over 2 weeks), we detected the same pattern in the two experiments (table A3). Overall, our results suggest that the between-status differences reflect genetic changes and, therefore, evolution of both species after invasion, rather than maternal effects. Conclusion Bioinvasions have long been considered interesting models for studying ecological and evolutionary processes (Baker and Stebbing 1965), but quantitative genetic approaches are still rare (e.g., Lee 2002; Facon et al. 2006). The emergence of novel interactions (trophic, competitive, or mutualistic) between native and invasive species is a potentially major source of change in selection pressures. We here showed that the emergence of a new competitive interaction due to invasion could trigger rapid evolutionary responses, as previously shown by Stuart et al. (2014). In addition, by studying both native and invasive species, we provide evidence for asymmetrical responses reflecting the different natures of selection pressures and constraints on them. Our data suggest that life-history trait shifts were large in the native species and tended to minimize the negative impacts of competition, while the evolutionary processes that took place in the invasive species during its spread were better explained by colonization dynamics and not necessarily adaptive. Finally, our study highlights the potential of life-history traits to underlie rapid evolutionary responses to new species interactions, in agreement with their often abundant genetic variance (Houle 1992) and their importance in models of species coexistence in metacommunities (Pavoine et al. 2014). More multispecies studies of this kind, applying evolutionary thinking to interacting invasive and native species in parallel, should be done to evaluate the generality of our hypotheses and to understand how communities change, not only ecologically but also evolutionarily, following invasions. Acknowledgments We thank G. Huth, H. Jourdan, and B. Pélissié for help running the experiment, the numerous people who participated in the metacommunity sampling, and C. Violle for discussions. E.C. was supported by an AXA postdoctoral fellowship. The study was made possible by a Chercheurs d’Avenir grant from the Région Languedoc-Roussillon and by an Agence Nationale de la Recherche grant (AFFAIRS, ANR12SV005) to P.D., as well as by support from CNRS to P.D. and P.J. This work is part of the COREIDS project supported by the Centre d’Analyse et Synthèse sur la Biodiversité and the Fondation pour la Recherche sur la Biodiversité. The authors would like to thank the three anonymous reviewers as well as the associate editor, M. Leibold, and the editor, A. Winn, for their valuable comments.
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Associate Editor: Mathew A. Leibold Editor: Alice A. Winn
