BIODIVERSITYRESEARCH

Going beyond taxonomic diversity:deconstructing biodiversity patternsreveals the true cost of iceplant invasionTommaso Jucker1, Marta Carboni2* and Alicia T. R. Acosta2

1Department of Plant Sciences, University of

Cambridge, Downing Street, Cambridge CB2

3EA, UK, 2Dipartimento di Biologia

Ambientale, Universit�a degli Studi di Roma

Tre, V.le Marconi 446, Roma 00146, Italy

*Correspondence: Marta Carboni,

Dipartimento di Biologia Ambientale,

Universit�a degli Studi di Roma Tre, V.le

Marconi 446, 00146 Roma, Italy.

E-mail: marta.carboni@uniroma3.it

ABSTRACT

Aim Although invasion has been linked to species losses in native plant com-

munities, it is unclear how invasive species affect other important aspects of

native community biodiversity, such as the composition of functional traits or

the degree of phylogenetic relatedness. Here, we ask whether declines in taxo-

nomic diversity (TD) associated with the spread of a highly invasive South

African species (iceplant) are linked to similar losses in functional (FD) and

phylogenetic diversity (PD). Motivated by recent advances in coexistence the-

ory, we aim to infer the mechanisms involved in driving the exclusion of native

species following invasion.

Location Coastal dunes of central Italy.

Methods We sampled 2 9 2 m vegetation plots characterized by varying

degrees of abundance of iceplant (Carpobrotus spp.), and combined species

occurrence data with life-history trait information and a dated phylogeny of

the native species pool. Rao’s quadratic entropy index was used to quantify the

TD, FD and PD of each plot, which we then related to iceplant abundance via

linear models. Finally, to better understand the mechanisms driving changes in

diversity, we characterized both species and communities according to phyloge-

netic relatedness to iceplant and functional strategy.

Results We found that the negative association between the level of invasion

and native community TD is mirrored by quantitatively similar declines in FD,

as well as a shift in community phylogenetic structure. These changes appear to

result from the selective exclusion of specific functional groups/clades, likely via

a combination of niche- and fitness-related processes.

Main conclusions We found that iceplant poses a greater threat to biodiversity

than previously understood and that on top of causing declines in species rich-

ness invasion may also hamper ecosystem functioning and reduce evolutionary

potential. Accounting for FD and PD holds promise for gaining a better under-

standing of how invasive species alter the structure of native communities.

Keywords

Biological invasions, Carpobrotus spp., coastal dunes, competitive exclusion,

fitness hierarchy, functional and phylogenetic diversity, invasive species.

INTRODUCTION

Invasive species constitute a major threat to biodiversity, and

their impact on natural and semi-natural habitats is the

focus of much ecological research (Pimentel et al., 2001;

Gurevitch & Padilla, 2004; Ehrenfeld, 2010). Numerous studies

have explored how invasion affects patterns of native species

diversity (Davis, 2003; Vil�a et al., 2011; Py�sek et al., 2012).

Yet despite the advent of readily accessible molecular phylog-

enies and comprehensive trait databases which have helped

redefine how we quantify and study biodiversity (Purvis

& Hector, 2000; Pavoine & Bonsall, 2011), most studies

DOI: 10.1111/ddi.121241566 http://wileyonlinelibrary.com/journal/ddi ª 2013 John Wiley & Sons Ltd

Diversity and Distributions, (Diversity Distrib.) (2013) 19, 1566–1577A

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continue to focus exclusively on how the number and rela-

tive abundance of native species (hereafter taxonomic diver-

sity, TD) is affected by invasion (Gaertner et al., 2009; Hejda

et al., 2009; Powell et al., 2011). In contrast, the impact of

invasives on other aspects of biodiversity such as functional

traits or phylogenetic relatedness remains unclear (e.g. Lamb-

don et al., 2008; Winter et al., 2009). Here, we investigate

how an invasive affects not only the TD of the native plant

communities it invades, but also their functional (FD) and

phylogenetic diversity (PD).

FD summarizes the diversity and variability of morpho-

logical, physiological and ecological traits present within a

community (Petchey & Gaston, 2006). The idea behind this

approach is to measure those traits which best capture a

species ecological niche. PD on the other hand is a measure

of the distance which separates each member of a commu-

nity when placed on a phylogeny (Purvis & Hector, 2000).

Assuming that closely related species are more functionally

similar than distant ones (Felsenstein, 1985; ‘phylogenetic

signal’ sensu Blomberg et al., 2001), the phylogenetic disper-

sion of a community should mirror its functional spectrum

(Cavender-Bares et al., 2004; Kraft et al., 2007; Vamosi

et al., 2009). Both the combination of functional traits and

the phylogenetic structure of a community have been shown

to play a key role in many ecological processes, such as

mediating species interactions and influencing community

assembly (Webb et al., 2002; HilleRisLambers et al., 2012),

determining responses to environmental change (Thuiller

et al., 2008; Swab et al., 2012) and driving ecosystem func-

tioning (Cadotte et al., 2011; Flynn et al., 2011). Conse-

quently, FD and PD are widely recognized as important and

complementary components of biodiversity and should be

studied alongside TD to better understand which factors

affect the structure and dynamics of plant communities

(Devictor et al., 2010), including responses to invasive

species.

Considerable effort has gone into understanding whether

certain traits make it more likely for species to become inva-

sive, as well as exploring how biodiversity (sensu lato) relates

to invasibility (Cadotte et al., 2009; MacDougall et al., 2009;

van Kleunen et al., 2010; Drenovsky et al., 2012). Less clear

is whether changes in competitive environment following

invasion, which have been linked to declines in TD, are also

altering the functional and phylogenetic structure of native

communities. Gerhold et al. (2011) reported that invasion

led to increased phylogenetic dispersion in recipient commu-

nities because of coexistence with, rather than replacement of

natives. It also appears that certain plant functional groups

may differ in their response to invasives (Davies, 2011). For

instance, Prunus serotina was shown to reduce FD in native

understorey plant communities by facilitating species with

specific trait combinations (Chabrerie et al., 2010). Overall,

these results suggest that focusing on multiple aspects of bio-

diversity can help us gain a more mechanistic understanding

of how invasives alter the structure of native plant

communities.

Determining how invasion might affect biodiversity pat-

terns requires an understanding of the processes which shape

FD and PD in plant communities. Traditionally, the

consensus has been that local scale diversity patterns essen-

tially depend on the combined effect of environmental filter-

ing (leading to the coexistence of ecologically similar

species), biotic interactions (leading to the exclusion of simi-

lar species, in accordance with niche theory) and neutral

processes related to dispersal and stochastic events (leading

to randomly assembled communities) (Hubbell, 2001; Webb

et al., 2002; Watkins & Wilson, 2003; Cavender-Bares et al.,

2004; Kraft et al., 2007). However, in the light of theoretical

advances in coexistence theory, this model of community

assembly has recently come into question (Mayfield &

Levine, 2010; HilleRisLambers et al., 2012). Crucially, as

Mayfield & Levine (2010) suggest, competitive interactions

often lead to the co-occurrence of functionally similar species

due to the exclusion of weaker competitors. To complicate

matters, further a number of other extrinsic factors also

affect local biodiversity patterns (e.g. spatial scales, regional

species pools, stress and disturbance; Pavoine & Bonsall,

2011; Brunbjerg et al., 2012; Mouillot et al., 2012). How the

disturbance caused by invasion affects the functional and

phylogenetic structure of communities is unknown.

We use Mediterranean coastal dune plant communities as

a model system to explore how the degree of invasion by the

South African succulent iceplant (Carpobrotus spp.) influ-

ences multiple aspects of native community biodiversity. Ice-

plant is considered a serious threat to biodiversity (Affre

et al., 2010) and has been associated with changes in coastal

ecosystems at a variety of scales (Vil�a et al., 2006; Conser &

Connor, 2009; Santoro et al., 2012b). Specifically, the pres-

ence of iceplant has been associated with a loss of TD in

native communities (Vil�a et al., 2006; Santoro et al., 2012a),

suggesting that the successful establishment of iceplant oper-

ates through the replacement and exclusion of incumbent

native species, rather than through coexistence with them.

However, it remains to be seen whether iceplant affects cer-

tain functional groups or phylogenetic lineages more than

others, thus resulting in altered patterns of FD and PD, or if

the loss of native species is generalized and independent of a

species ecological role. Assuming that iceplant invasion leads

to a decrease in TD (i.e. loss of species, as evidenced by pre-

vious studies), two alternative mechanisms of species exclu-

sion can be hypothesized, each leading to distinct FD and

PD patterns (Fig. 1):

H1)Generalized/Random impact: the loss of TD is not

associated with a decrease in FD and PD, which remain

unchanged. According to this scenario, iceplant randomly

excludes species from the community, independently of

their functional characteristics or phylogenetic clade. Ice-

plant competes equally with all species in the community,

probably for basic resources such as space, light or water.

H2)Directional/Selective impact: the decrease in TD is

coupled with a parallel decrease in FD and/or PD, as ice-

plant selectively excludes species that share a specific set

Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd 1567

Invasion and biodiversity loss in plant communities

of functional traits or belong to a particular lineage. In

this scenario, the presence of iceplant could either lead to

the loss of species that share its ecological requirements

(limiting similarity hypothesis; H2a) or to the exclusion

of those that are competitively inferior (weaker

competitor exclusion hypothesis; H2b).

METHODS

Study area and vegetation sampling

We studied recent (Holocene) coastal dunes of central Italy

(Lazio Region) at six sites distributed along the length of

Figure 1 Conceptual diagram illustrating how alternative processes of species exclusion in response to invasion (left-hand side) can

lead to contrasting diversity patterns (right-hand side). Communities are characterized both in terms of their phylogenetic structure and

by a combination of two quantitative traits (symbolized by squares and triangles, which vary in sizes according to unitless values).

Following invasion, extant lineages are represented by filled symbols (black squares and triangles), while excluded species are shown as

empty symbols. According to the first hypothesis (H1; top row), invasion leads to the exclusion of a random subset of species (both in

terms of their position on the phylogeny and of their trait values). As a result, although TD declines in response to invasion, as long as

PD and FD are not correlated with TD, they both remain unaffected by the processes. Conversely, in the second scenario, the

introduction of an exotic species triggers the loss of specific lineages and traits from the species pool, leading to a decline in both PD

and FD. However, this same pattern can result from two very different processes. In the first instance (H2a), in accordance with the

limiting similarity hypothesis, the invader competes most strongly with those species that share its ecological requirements (in the

diagram, excluded species have a high degree of niche overlap with the invader in terms of their trait values). Instead, based on the

predictions of the weaker competitor exclusion hypothesis (H2b), the change in competitive regime brought on by invasion should lead

to the loss of competitively inferior species (here symbolized by smaller trait values).

1568 Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd

T. Jucker et al.

coastline (Fig. 2). The area is characterized by a Mediterra-

nean climate, with recent dunes generally occupying a nar-

row strip along the seashore. The dunes are not high

(<10 m) and are relatively simple in structure, with beaches

varying in breadth from a few metres to around 40 m, fol-

lowed by a section of low embryo-dunes, generally only one

mobile dune ridge, dune slacks and lastly a stabilized dune

zone (Acosta et al., 2003; Carboni et al., 2011). Vegetation

sampling was undertaken in spring (April–May) between

2006 and 2009. Randomly generated GPS coordinates were

used to define the location of 2 9 2 m vegetation plots

which we then sampled in the field. In each plot, we

recorded the presence of all vascular plant species (natives

and aliens), along with a measure of percentage cover for

each species using a 10% interval rank scale (for further

details on sampling and nomenclature, see Santoro et al.,

2012a).

A total of 414 plots were visited and 158 species recorded.

Based on previous work which identified six main vegetation

habitats along the sea-inland zonation (from pioneer com-

munities of the upper beach to shrubby communities in the

back dune; Carboni et al., 2011), each study plot was

assigned a habitat code (coding scheme follows that of the

European Habitats Directive; European Commission 1992).

Iceplant invasion occurs almost exclusively in the three habi-

tats which occupy the central sector of the zonation (EU

habitats 2110, 2210 and 2250; Carboni et al., 2011), which

are characterized by comparable levels of iceplant abundance

(habitat 2110: 37.7 � 4.5%; habitat 2210: 32.2 � 5.2%; habi-

tat 2250: 39.3 � 8.4%). For the purpose of this study, we

therefore restricted our analyses to plots found in this section

of the dunes. Furthermore, in order to test the extent to

which plot-level patterns of TD, FD and PD are affected by

the degree of iceplant invasion (quantified in terms of

percentage cover), we focused primarily on plots in which

iceplant was recorded (n = 58). Because exploring what

drives biodiversity patterns in native communities goes

beyond the focus of this study, data from non-invaded plots

were only used to provide context for certain results and

were excluded from further analyses unless explicitly stated.

Plant functional traits

Traits were compiled and measured for a subset of 46 domi-

nant species (Ricotta et al., 2012). Collectively, these selected

species account for ~80% of the standing live biomass in

each dune habitat. Seven life-history traits were selected

based on their relevance to the functional ecology of plants

in coastal dune environments: plant height (cm), SLA

(mm2 mg�1), LDMC, leaf size (cm2), life-form, dispersal

mode and pollination system. Quantitative traits (plant

height, SLA, LDMC, leaf size) were measured in the field or

laboratory by taking the mean of 10 different individuals.

Qualitative traits (life-form, dispersal mode, pollination sys-

tem) were obtained from the literature and were supple-

mented by personal observations in the field (for details see

Ricotta et al., 2012).

Phylogeny and phylogenetic signal

Using the complete species list, we constructed an aged phy-

logenetic tree using the open source software Phylomatic and

Phylocom (Webb et al., 2008). Branch lengths were assigned

using a branch length adjustment algorithm (BLADJ), based

on the minimum age of nodes estimated from the

fossil record. We then used a standard Mantel statistics

(Hardy & Pavoine, 2012) to test whether the multivariate

matrix of pair-wise functional dissimilarities between species

Figure 2 Study area. Locations where

vegetation sampling was undertaken are

marked by open squares.

Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd 1569

Invasion and biodiversity loss in plant communities

(calculated with the Gower distance for mixed-variables) was

significantly correlated with the matrix of phylogenetic dis-

tances between species (summed branch lengths separating

pairs of species). Gower was chosen as it summarizes the

information of multiple traits (both quantitative and qualita-

tive) into a single metric, which for the purposes of this

study was preferable to analysing each trait individually. We

found a significant Mantel correlation (R = 0.215;

P = 0.001) between species’ pair-wise functional distances

and their phylogenetic relatedness, meaning that the traits we

selected exhibit strong phylogenetic signal at the species

level.

Quantifying taxonomic, functional and phylogenetic

diversity

Among the numerous metrics that exist to quantify biodiver-

sity patterns (Pavoine & Bonsall, 2011), we selected Rao’s

quadratic entropy index as it provides a framework for cal-

culating different biodiversity components (i.e. TD, FD and

PD) using a common methodological approach (de Bello

et al., 2010). This allows a much more robust comparison

between the different diversity measures. For a given plot, we

calculated Rao’s alpha diversity as follows:

aRao ¼Xs

i¼1

Xs

j¼1

dijpipj

where s is the number of species in the plot (species rich-

ness), dij is the distance between the species pair i and j,

weighted by the relative abundance pi and pj of the two spe-

cies (estimated here on the basis of percentage cover in the

plot). For FD and PD, the dissimilarity between species (dij)

is quantified using the multivariate functional distance

matrix (using the Gower distance) and the phylogenetic dis-

tance matrix, respectively (de Bello et al., 2010). In the case

of TD, where the distance between two species is fixed

(dij = 1, unless i = j in which case dij = 0), a Rao is equiva-

lent to the Simpson diversity index. All calculations were

performed in R (2.15; R Foundation for Statistical Comput-

ing, Vienna, Austria. http://www.R-project.org.) by adapting

code provided in the supplementary material of de Bello

et al. (2010).

Accounting for collinearity among diversity metrics

As TD, FD and PD are derived from a common species

abundance matrix, they will inevitably exhibit a certain

degree of correlation between one another. Unsurprisingly,

Spearman’s rank correlation coefficients (q) among the three

diversity indices were positive (qTD:FD = 0.73; qTD:PD = 0.61;

qPD:FD = 0.37) and highly significant (P < 0.001; Fig. 3).

This strong collinearity poses a problem when attempting to

infer the mechanisms through which iceplant excludes native

species from a community (directional vs. generalized),

because any decline in TD will lead to a decrease in FD and

PD even if species loss is random. To account for this, we

first regressed FD and PD against TD (all logit-transformed)

and then adopted the residuals of these regressions as

response variables in our models (FDRES and PDRES). By

determining what portion of the residual variance is

explained by iceplant abundance, we were thus able to infer

the direct impact of invasion on a community’s functional

and phylogenetic structure (Devictor et al., 2010). As an

alternative approach, we accounted for variation in TD using

a null model that randomizes functional and phylogenetic

relationships among species to standardize FD and PD values

(see Appendix S1 in Supporting Information). Given that

both methods yielded almost identical results (Fig. S1), we

focus on the simpler regression residuals approach here.

Modelling TD, FD and PD as a function of iceplant

abundance

Linear models were fit for each of the three response vari-

ables (TD which we logit-transformed, FDRES and PDRES).

Iceplant percentage cover was adopted as the main explana-

tory variable. Alongside this, we accounted for the habitat

type of each plot to accommodate the fact that native plant

diversity differs markedly between habitats (hereafter referred

to using their respective European Habitats Directive codes:

2110, 2210 and 2250; Acosta et al., 2009). Importantly, an

interaction term between iceplant cover and habitat was also

fit to determine whether invasion impacts the three habitats

differently. Lastly, three further covariates were included in

the models to correct for additional sources of variation in

diversity patterns. A measure of distance from the sea of

each plot was calculated in a GIS environment (ArcGIS 9.2,

ESRI, Redlands, CA, USA) and used as a proxy of environ-

mental harshness (salt spray, sand burial, soil aridity, salinity

and nutrient shortage are all known to decrease with distance

from the sea; Carboni et al., 2011). Similarly, the distance

from the nearest human structure was also computed and

adopted as an indicator of human disturbance (Alston &

Richardson, 2006; Carboni et al., 2011). Finally, the abun-

dance of all other alien plants was used to account for the

potential influence of other invasive species. Model simplifi-

cation was performed by implementing stepwise model selec-

tion (forward and backward) based on AIC.

Determining which phylogenetic lineages and

functional traits are targeted by iceplant

Two complementary approaches were taken to determine

whether species being excluded by iceplant are more likely

to be functionally and phylogenetically similar (in accor-

dance with the limiting similarity hypothesis) or different

from the invader (suggesting the weaker competitor exclu-

sion hypothesis): one exploring individual species responses

and the other focusing on community-level patterns.

Assuming that iceplant selectively excludes species from a

community, it should be possible to quantify the sensitivity

1570 Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd

T. Jucker et al.

of individual species to iceplant and then explore which

traits best predict how species respond to invasion. Follow-

ing the same approach used to infer potential impacts of

iceplant on community-level diversity metrics, we therefore

fit separate linear models relating individual species abun-

dances to iceplant cover for each of the 46 dominant spe-

cies for which trait data were available. Again, habitat type,

distance from the sea and human structures, and abundance

of other invasives in each plot were adopted as covariates in

the models. In addition, we also accounted for cumulative

vegetation cover in each plot. This allowed us to control for

the fact that the abundance of individual species is expected

to decline in plots with higher vegetation cover as a result

of decreased available growing space. For each species, we

then extracted the slope parameter (b) for the term relating

to iceplant cover from the models and used it as a means

of quantifying a species’ sensitivity to invasion: the more

negative b is, the greater the negative association between

iceplant and the target species. Finally, we used linear mod-

els to relate a species’ sensitivity to invasion to its func-

tional trait portfolio in order to determine whether specific

life-history traits make species more susceptible to competi-

tive exclusion by iceplant. We also tested whether the likeli-

hood of a species decreasing in abundance in response to

invasion was associated with its rarity, quantified in terms

of both its mean cover and presence frequency in all non-

invaded plots. If species losses are random and not medi-

ated by niche/fitness processes (H1), then we might expect

rare species to have a greater probability of being excluded

from a community through chance alone (Suding et al.,

2005).

To understand how the phylogenetic structure of native

communities responds to invasion, we explored a series of

metrics which quantify the degree of evolutionary relatedness

between an invader (iceplant) and other members of a com-

munity (Thuiller et al., 2010). Specifically, we computed the

Mean Distance of the introduced species relative to that of

Native Species (MDNS) and the Weighted Mean Distance to

the Native Species (WMDNS) using the phylogenetic dissim-

ilarity matrix and the data on community composition. We

then quantified how MDNS/WMDNS varied in relation to

the degree of invasion using linear and quantile regressions

(package ‘quantreg’ in R). An increase in MDNS in heavily

invaded plots would highlight the exclusion of phylogeneti-

cally close (and presumably functionally similar) species,

whereas a decrease in MDNS would suggest the loss of dis-

tantly related species. To compare invaded and non-invaded

plots and test the robustness of observed patterns, we also

computed ‘theoretical’ MDNS values for sites where iceplant

was not present.

RESULTS

Our analyses revealed a negative trend between iceplant inva-

sion and all three diversity metrics being examined. How-

ever, this association was only significant in the case of TD

and FD, while the observed decline in PD was not robustly

supported in our models (Fig. 4). Our results also suggest

that across all diversity components, iceplant is eliciting

greater negative responses in habitats closer to the sea (habi-

tat 2110) compared to communities found further along the

zonation (habitat 2250).

Taxonomic diversity

A significantly negative relationship between iceplant cover

and native species TD emerged from the model (P = 0.005;

slope parameter = �0.17 � 0.06). This pattern of decreasing

TD in relation to the increased abundance of iceplant per-

sisted even after having accounted for the confounding

effects of habitat type, distance from the sea, distance from

human structures and the abundance of other alien species.

In fact, model simplification led to the exclusion of all cova-

riates except habitat type and the interaction term between

iceplant abundance and habitat. Unsurprisingly, TD

increased along the habitat zonation. Instead the interaction

term revealed that iceplant had the greatest impact on TD in

habitat 2110 which is closest to the sea (P < 0.001; slope

parameter = �0.30 � 0.09). The final model explained

~35% of the total variance (R2 = 0.33).

Figure 3 Scatterplots illustrating the correlation among (a) TD–FD, (b) TD–PD and (c) PD–FD. Spearman’s rank correlation

coefficients (q) are printed in the top-left of each panel.

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Invasion and biodiversity loss in plant communities

Functional diversity

Iceplant abundance also played an important role in shaping

FD patterns, even after having accounted for the strong cor-

relation between FD and TD. In general, the minimal model

explained ca. 25% of the variance (R2 = 0.24) in FDRES val-

ues, and iceplant cover was associated with a significant

decrease in native species FD (P = 0.022; slope parame-

ter = �0.14 � 0.06). Habitat type was again the only covari-

ate to be retained in the minimal model, and as was the case

for TD, the interaction term suggests that the impact of ice-

plant on FDRES decreases along the zonation (becoming non-

significant in habitat 2250; slope parameter = 0.03 � 0.08).

Phylogenetic diversity

Although PD declined markedly in response to invasion,

once the strong correlation with TD was accounted for, ice-

plant abundance explained only a marginal portion of the

residual variation in PDRES (P > 0.05; slope parame-

ter = �0.06 � 0.06). Model simplification led to the exclu-

sion of all predictors aside from habitat type, which alone

accounted for 57% of the variance in PDRES. This very

strong effect was primarily driven by the sharp increase in

PD which occurs in habitat 2250 where several outgroup lin-

eages enter the species pool. Although the effect of the inter-

action term was non-significant, it is interesting to note that

PDRES was also more susceptible to declines in habitat 2110.

Life-history traits as predictors of individual species

responses to invasion

Consistently with community-level analyses, individual spe-

cies models revealed that, on average, species’ abundances

tended to be negatively related to iceplant cover

(b = �0.05 � 0.05). Nevertheless, susceptibility to invasion

varied considerably among species, largely as a result of dif-

ferences in life-history strategies. Three traits in particular

(SLA, plant height and seed dispersal mode) emerged as

strong indicators of whether a species would be negatively

related to iceplant and together explained 43% of the varia-

tion in individual species responses to invasion (Fig. 5). Spe-

cies with low SLA were much more likely to decline when

iceplant was abundant (P < 0.001). Similarly, low-growing

plants (maximum height <50 cm) also tended to respond

negatively to invasion, while those attaining a height of more

than 50 cm were largely unaffected (P = 0.071). Lastly, spe-

cies that rely on wind for seed dispersal were much more

Figure 4 Change in Taxonomic (TD, logit-transformed) and

Functional (FDRES) diversity in relation to the degree of

invasion by iceplant (measured in terms of% ground cover).

Plotted lines with shaded confidence intervals represent the fit of

the minimal adequate models for both TD (solid dark grey line)

and FDRES (dotted light grey line).

(a) (b) (c)

Figure 5 Plots showing the relationship between individual species responses to invasion (b) and life-history traits values for (a) SLA,

(b) plant height and (c) seed dispersal mode. For comparison purposes, iceplant trait values are as follows: SLA = 5.53 mm2 mg�1;

height = 19.8 cm; animal dispersed. P-values are printed on the bottom-right of panel A and B, while in panel C, letters refer to

independent groups according to Tukey HSD test.

1572 Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd

T. Jucker et al.

strongly impacted than those that disperse through other

abiotic mechanisms (ballistic, water) or are spread by ani-

mals. In contrast to life-history traits, we found no evidence

to suggest that susceptibility to iceplant is associated with the

commonness or rarity of a species (Fig. S2).

Changes in community phylogenetic structure

following invasion

MDNS increased significantly (P = 0.026; slope parame-

ter = 0.084 � 0.038) in response to invasion (Fig. 6), with

MDNS values being lowest in non-invaded plots (the same

pattern was also found for WMDNS values; Fig. S3). This

trend could result either from the loss of closely related taxa

or colonization by phylogenetically distant species. However,

quantile regression on the lower boundary of the data

(s = 0.1) revealed that this shift in mean was primarily dri-

ven by a loss of communities with low MDNS values

(P < 0.001; slope parameter = 0.097 � 0.028), suggesting

that iceplant preferentially excludes closely related species.

Together with our results showing that functional traits help

predict whether a species will be negatively related to iceplant

abundance, our analysis of community phylogenetic structure

suggests that species losses following iceplant invasion are

more selective than generalized (H2).

DISCUSSION

Accounting for life-history traits and phylogenetic history is

proving extremely useful in the search for general principles

that help explain why certain plant species become invasive

(Hayes & Barry, 2008; Cadotte et al., 2009; van Kleunen

et al., 2010; Drenovsky et al., 2012). Yet in terms of quanti-

fying how invasives impact native plant communities, most

studies continue to focus primarily on how species richness

is affected by invasion. Here, we show that previously docu-

mented decreases in TD following the establishment of ice-

plant in coastal dune ecosystems (Vil�a et al., 2006; Santoro

et al., 2012a) are coupled with comparable declines in FD, as

well as being associated with a shift in the phylogenetic

structure of native communities.

In the introduction, we hypothesized two processes

through which iceplant could competitively exclude native

species from a community and suggested that the two mech-

anisms would give rise to distinct FD and PD patterns. Spe-

cifically, we reasoned that if iceplant were to compete equally

with all species (e.g. neutral theory), then species losses

would be random and would not necessarily lead to a

decrease in FD or PD (H1). If instead iceplant preferentially

excludes certain clades or functional groups, then a loss in

TD would most likely be accompanied by a concurrent

decline in FD and PD (H2). Central to both of these hypoth-

eses is the assumption that iceplant is indeed associated with

a decline in TD, as previous studies suggest (Vil�a et al.,

2006; Santoro et al., 2012a). As expected, iceplant abundance

was found to correlate negatively with TD in our study sites.

This finding is consistent with much of the literature regard-

ing how exotic plants affect the communities they invade

(Gaertner et al., 2009).

In conjunction with the decline in TD, a parallel loss in

FD was also observed as iceplant abundance increased

(Fig. 4). This pattern is consistent with the predictions of

H2, according to which certain functional groups and/or

phylogenetic lineages should be more susceptible to invasion

than others. This suggests that iceplant is somehow acting as

a filter in the process of community assembly by preferen-

tially excluding species with specific life-history traits. The

selective exclusion of a given trait can result from direct

competition with the invader for resources (e.g. for space,

water or nutrients). Iceplant is known to strongly compete

with native species for space and water (Albert, 1995), lead-

ing to the reduced growth, reproductive output and survival

of neighbouring plants (D’Antonio & Mahall, 1991; Conser

& Connor, 2009). Alternatively, exclusion can occur indi-

rectly as a consequence of decreased habitat suitability fol-

lowing the establishment of the novel species. Iceplant has

been shown to drive significant changes in soil chemistry

(e.g. [C], [N], pH), which are likely to affect the fitness of

certain functional groups more than others (Vil�a et al., 2006;

Santoro et al., 2011). Although we observed a shift in phylo-

genetic structure in invaded communities, we found no

direct influence of iceplant on PD. The strong correlation

Figure 6 Scatterplot showing the relationship between Mean

Distance to Native Species (MDNS; log-transformed) and

iceplant vegetation cover (%). Circles represent mean MDNS

values for each level of iceplant cover, with bars corresponding

to standard errors. The solid line shows the fit of a linear model.

Non-invaded plots (left of dotted line) are shown as an empty

circle, while filled circles represent mean values for invaded

plots. The standard error for non-invaded plots was calculated

on a random subsample of 30 sites and then averaged across

1000 iterations. This was carried out to provide comparable

estimates of uncertainty between invaded (58) and non-invaded

plots (191).

Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd 1573

Invasion and biodiversity loss in plant communities

between PD and TD could partly account for this, because

little residual variation in PD remained for iceplant abun-

dance to explain. A better resolved molecular phylogeny

could potentially provide the phylogenetic precision needed

to detect shifts in PD.

Aside from revealing that iceplant influences more than just

TD, our results also hint to which processes are driving the

competitive exclusion of species. For much of the last decade,

the prevalent theory has been that regional species pools are

constrained by habitat filtering, while at the community-level

coexistence, patterns can be explained by the limiting similar-

ity hypothesis (H2a), which predicts that ecologically similar

species will exclude one another due to strong niche overlap

(Webb et al., 2002; Emerson & Gillespie, 2008). However, an

alternative model based on coexistence theory (Chesson,

2000) attempts to explain why species co-occurrence patterns

often clash with traditional expectations (HilleRisLambers

et al., 2012). This framework proposes that the outcome of

competition among species depends on the interplay between

niche differences (ecological similarity) and relative fitness dif-

ferences (competitive ability), thus explaining why it is often

the competitively weaker species that are excluded (H2b). Our

analyses suggest that changes in the functional and phyloge-

netic structure of native communities associated with high

levels of iceplant invasion may be the result of two separate

processes. In line with the predictions of limiting similarity

hypothesis (H2a), iceplant appears to preferentially exclude

closely related and ecologically similar taxa (Figs 5 & 6). Low-

growing species with small leaf surface area-to-weight ratios

(similar to iceplant: height = 19.8 cm; SLA = 5.53 mm2

mg�1) were much more likely to decrease in abundance in

response to invasion. In contrast, species that grow above ice-

plant mats and/or have high SLA were largely unaffected, pos-

sibly as a result of their ability to avoid excessive shading

(Weiner & Thomas, 1986; Lep�s, 1999).

However, iceplant also had a strong impact on species that

are wind dispersed. Limited seed reserves may make it harder

for wind dispersed species to germinate and grow under the

suboptimal conditions associated with iceplant mats (e.g.

altered soil chemistry, reduced water and light availability;

D’Antonio & Mahall, 1991; Vil�a et al., 2006; Conser & Con-

nor, 2009; Santoro et al., 2011). The fact that iceplant may

be limiting the distribution of wind dispersed species, many

of which are ephemeral and characterized by annual life

cycles, suggests that weaker competitors are also being

excluded from the species pool (H2b). This is consistent with

previous studies showing that annual grasses (therophytes)

are particularly sensitive to iceplant (Vil�a et al., 2006; Andreu

et al., 2010), a pattern which we also find (Fig. S4). More-

over, it is supported by the fact that iceplant had the strong-

est impact in foredune communities where many species

have annual life cycles (H2b) and/or adopt creeping growth

strategies (H2a). Although seemingly contradictory, the fact

that both weaker competitors and ecologically similar species

declined in the presence of iceplant is precisely what

coexistence theory would predict. Not only are niche- and

fitness-related processes non-mutually exclusive, it is their

combined effect which ultimately shapes community

assembly and species exclusion following invasion.

As with any correlative study, caution is needed when inter-

preting results (Renter�ıa et al., 2012). Because we have no

information on the floral composition of invaded plots prior

to invasion, it is impossible to determine whether the lower

diversity of these plots is a direct consequence of invasion or

simply a pre-existing pattern which facilitated iceplant estab-

lishment (MacDougall & Turkington, 2005; Carboni et al.,

2013). Nevertheless, we feel there are several reasons why our

interpretation of the data is well supported by our analysis.

First, ours is not the first study to identify a change in commu-

nity structure and diversity following invasion by iceplant,

even though previous work had not explored FD and PD pat-

terns. Furthermore, because our analysis relates the loss of

diversity to the relative abundance of iceplant (and not to its

presence/absence), we were able to show that TD and FD pro-

gressively decline as iceplant becomes more abundant. If ice-

plant is not driving species losses and is merely a passenger

along for the ride (MacDougall & Turkington, 2005), we might

instead expect little difference in diversity among weakly and

strongly invaded plots. Finally, iceplant abundance was nega-

tively associated with TD and FD even after having accounted

for environmental stress and human disturbance, suggesting

the patterns we find are not driven by external factors. Regard-

less, future experimental and observational work is needed to

settle this debate. Long-term monitoring, both in the field and

under controlled experimental settings, would allow changes

in community composition to be assessed over time. Con-

versely, removal trials could help determine whether currently

invaded plots shift back to baseline conditions, or if these com-

munities were in fact compositionally distinct to start with.

The need to redefine conservation priorities and adapt

management strategies to include aspects such as functional

and phylogenetic diversity is clear (Mace et al., 2003; Forest

et al., 2007; Cadotte et al., 2011; Rolland et al., 2012). Real-

izing that invasive species may affect not only patterns of

species richness, but also hamper ecosystem functioning and

reduce evolutionary potential is an important step in this

process. Conservation efforts assessing the impact of inva-

sives need to consider that native communities can be

affected in more ways than one, or underlying changes could

be inadvertently missed. Just as importantly, in order to

develop effective management strategies to safeguard biodi-

versity, the knowledge that some species may be more threa-

tened than others by invasion is crucial. In the hope of

preserving the functionality and uniqueness of coastal dune

ecosystems, mitigating the spread of exotic species such as

iceplant should be of special concern.

ACKNOWLEDGEMENTS

We are grateful to R. Santoro for helping us develop and

refine the questions explored here. Three anonymous

reviewers, along with editors David Richardson and Mark

1574 Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd

T. Jucker et al.

van Kleunen, were instrumental in helping us improve on

earlier drafts. Lastly, we thank S. Del Vecchio and I. Prisco

for helping with fieldwork.

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SUPPORTING INFORMATION

Additional Supporting Information may be found in the

online version of this article:

Appendix S1 Comparison of ‘regression residuals’ and ‘null

model’ approaches to removing the correlation between FD

and PD values with TD.

Appendix S2 Test of the ‘random loss scenario’ according to

which rare or low-abundance species are more likely to be

excluded from a community through chance alone.

Appendix S3 Comparison of Mean Distance to Native Spe-

cies (MDNS) and Weighted Mean Distance to Native Species

(WMDNS) metrics used to quantify the mean phylogenetic

distance between iceplant and all other species present in a

community.

Figure S1 Scatterplot of regression residuals (RES) and stan-

dardized effect sizes (SES) for FD and PD.

Figure S2 Relationship between species sensitivity to invasion

and two metrics of species rarity - mean vegetation cover

and frequency of plots in which the species is found.

Figure S3 Scatterplot showing the relationship between

Weighted Mean Distance to Native Species (WMDNS) and

iceplant vegetation cover.

Figure S4 Comparison of the mean number of annual grass

species (Therophytes) found in plots with varying degrees of

iceplant abundance.

BIOSKETCHES

Tommaso Jucker is a plant ecologist interested in the

mechanisms involved in promoting, maintaining and shaping

biodiversity patterns. His research has explored how plant

communities respond to drivers of global change and more

recently has focused on understanding how biodiversity loss

influences ecosystem functioning. He is currently pursuing a

PhD at the University of Cambridge, before which he was a

visiting researcher at the University of Roma Tre, in the

plant ecology laboratory lead by Alicia T.R. Acosta. Her

research team relies on coastal dune environments as model

systems to analyse distribution patterns and dynamics of

plant species/communities in relation to environmental and

human drivers and to agents of global change (particularly

invasive species) across spatial scales. Marta Carboni is a

post-doctoral researcher focusing on plant invasions and on

the mechanisms driving community assembly and species co-

existence.

Author contributions: A.A., M.C. and T.J. conceived the

ideas; M.C. and T.J. collated and analysed the data; all

authors contributed to the writing which was led by T.J.

Editor: Mark van Kleunen

Diversity and Distributions, 19, 1566–1577, ª 2013 John Wiley & Sons Ltd 1577

Invasion and biodiversity loss in plant communities