Review
Introduced Species, Disease
Ecology, and Biodiversity–
Disease Relationships
1, 2 3
Hillary S. Young, * Ingrid M. Parker, Gregory S. Gilbert,
1 4,5
Ana Sofia Guerra, and Charles L. Nunn
Species introductions are a dominant component of biodiversity change but are
Trends
not explicitly included in most discussions of biodiversity–disease relationships.
Introduced species can have strong
This is a major oversight given the multitude of effects that introduced species effects on parasite prevalence and
richness.
have on both parasitism and native hosts. Drawing on both animal and plant
systems, we review the competing mechanistic pathways by which biological
Introductions primarily affect disease
fl
introductions in uence parasite diversity and prevalence. While some mecha- via changes in host composition, not
richness.
nisms – such as local changes in phylogenetic composition and global homog-
enization – have strong explanatory potential, the net effects of introduced
Local effects of introduced species can
species, especially at local scales, remain poorly understood. Integrative, com- amplify or dilute parasite prevalence.
munity-scale studies that explicitly incorporate introduced species are needed
Global homogenization and spread of
to make effective predictions about the effects of realistic biodiversity change human commensals systematically
and conservation action on disease. increase disease.
Global Change, Diversity and Disease
Anthropogenic disturbances are driving catastrophic rates of species decline and extinction [1].
Meanwhile, biological introductions are increasing rapidly [2,3]. Together this has resulted in
dramatic changes to the species diversity (see Glossary) and composition – the biodiversity –
in most ecosystems [4,5]. Such changes can directly impact many ecosystem services [6],
including disease spread and impact. Systematic effects of biodiversity change on infectious
disease could have enormous economic, public health, and conservation implications. As
examples, at least 60% of all human disease agents are zoonotic in origin [7], plant disease
is blamed for the loss of 10% of global food production [8], and infectious disease is among the 1
Department of Ecology, Evolution,
top five drivers of global extinction [9]. Appropriately, then, a robust literature has emerged on
and Marine Biology, University of
the likely relationship between biodiversity change and the prevalence of infectious disease California, Santa Barbara, Santa
– Barbara, CA, USA
[10 12]. 2
Department of Ecology and
Evolutionary Biology, University of
Much of this work has focused on understanding how anthropogenically driven losses in native California, Santa Cruz, CA, USA
3
Department of Environmental Studies,
species diversity affect transmission of parasites and the disease landscape [11,13]. Primary
University of California, Santa Cruz,
research and meta-analysis supporting a negative relationship between species richness and
CA, USA
4
parasite prevalence is often interpreted as evidence for the disease control value of conservation Department of Evolutionary
Anthropology, Duke University,
[11,12]. Often, this negative relationship between diversity and disease is attributed to the
Durham, NC, USA
dilution effect (Box 1) [11], in which hosts in high-diversity systems have lower average 5
Duke Global Health Institute, Duke
competence for a given parasite thus reducing transmission and community level prevalence University, Durham, NC, USA
for that parasite. If this pattern is general across multiple parasites, it produces an overall
negative correlation between host diversity and parasite prevalence. However, while there is
*Correspondence:
elegance in simple appeals for the conservation of natural biodiversity (or for increasing [email protected]
biodiversity in managed landscapes), the mechanistic relationships between biodiversity change (H.S. Young).
Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 http://dx.doi.org/10.1016/j.tree.2016.09.008 41
© 2016 Elsevier Ltd. All rights reserved.
Box 1. Effects of Species Richness and Composition on Disease Prevalence Glossary
Pathogens cause disease when they successfully infect a susceptible host. Some infected hosts are tolerant to infection
Abundance: when used in context
and suffer little impact of disease. All susceptible hosts (regardless of tolerance) can be competent to support
of parasites, this refers to the mean
reproduction of the parasite. When a population has enough hosts that are both susceptible and competent, an
parasite load of host individuals in the
infectious disease can spread through the population. In other cases, a susceptible host may be strongly affected by the
entire host population regardless of
disease but be a dead-end (non-competent) host for pathogen reproduction and transmission; that host would not
whether a host is infected.
contribute to additional disease pressure in the population.
Alpha diversity: mean species
diversity at local scales.
Disease spread can be density dependent (increased transmission at higher host densities for passively dispersed Amplification effect: when diverse
pathogens) or frequency dependent (increased transmission with greater frequency of encounter of susceptible host communities cause increased
individuals with infectious individuals or vectors). Any changes to community diversity or composition that reduce rates of parasite prevalence and
the likelihood that parasites will encounter susceptible hosts should reduce disease pressure. In a saturated community, transmission for a given parasite,
when species richness increases (Figure I, A C or B D), the average number of individuals per species generally often because of increases in the
) )
declines. While for true generalist parasites (for which all hosts are equally competent) such changes in host composition relative abundance of competent
will not matter, for host-specialized parasites this greater species richness can then ‘dilute’ competent hosts and reduce hosts (amplifying hosts).
disease pressure. Similarly, loss of species richness (e.g., conversion of a wild grassland to an agricultural monoculture; Amplifying hosts: competent hosts
Figure I, D B), can increase disease pressure. However, most parasites are not strict host specialists, so the overall with high transmission efficiency for
)
abundance of competent hosts in the community has the stronger effect on pathogen transmission; thus, whether that parasite.
increases in species richness are associated with increases in competent or non-competent hosts is crucial. For Biodiversity: a measure of the
instance, replacing half of the individuals in a monoculture of a susceptible host with one non-competent host species variety and variability of species of
(Figure I, A E) has a greater dilution effect than adding a dozen species that include competent hosts (Figure I, A C). organisms found in different
) )
However, a similar doubling of the richness of a non-competent monoculture to include two low-competence hosts ecosystems; it includes aspects of
(Figure I, B E) has little impact on disease pressure. Understanding the impacts of invasive species on disease pressure both composition and diversity and is
)
thus requires the evaluation not only of changes in species richness but also of changes in the composition, traits, and measured with a range of metrics
relative abundance of host species. including richness, evenness, and
phylogenetic diversity.
Coextinction: when loss of a
phylogenetic lineage leads to
A secondary loss of closely associated
lineages, which can include any type
of symbiotic relationship (parasitism,
mutualism, or commensalism).
C
Coextinction is more likely when the
associate is specialized on one or a D
i s
few lineages. e
a
s Competence: capacity of a host to e
p
become infected by a particular r
e
s parasite and make it available for s u
further transmission. r e E
B Density-dependent transmission: s
t
s transmission is a function of the
o
h density of the host population, as t
n
e might occur with a socially t
e
p transmitted parasite in a highly mixed
m
population or a passively dispersed Sp o e c ci D e plant pathogen. s r n
ich o
n Diluting hosts: hosts with low es r
s o
p competence for a given parasite and o
r
P thus low transmission efficiency.
Dilution effect: when more diverse
host communities cause reduced
rates of parasite prevalence and
Figure I. Example of Expected Effects of Host Species Richness and Composition on Parasite Richness
transmission, typically because of
and Prevalence.
reductions in the relative abundance
of competent hosts in diverse
communities for a given parasite;
similar to the decoy effect in
and disease are more complicated – and interesting – than parasite abundance responding to
macroparasite studies (Box 1).
simple declines in native host species richness.
Disease: represents the fitness cost
to host individuals that are infected
Remarkably, given that increases in the number of introduced species often numerically by parasites or microbial pathogens.
Disease-mediated invasion:
compensate for modern losses in native biodiversity [5,14–17], the importance of introduced
occurs when a parasite is shared (via
species has often been ignored in the discussion of biodiversity–disease relationships [11–13]
spillover or spillback) between native
(but see [18,19]). This omission is particularly surprising because introduced species are closely
42 Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1
associated with emergent diseases and can impact disease pressure in a community in and introduced host species and has
higher virulence in native species,
multiple ways [20]. Moreover, a recent meta-analysis of diversity–disease relationships [11]
facilitating the spread of the
found only three studies examining the effects of introduced species on infectious disease risk in
introduced species.
native ecosystems via changes in biodiversity; these studies produced inconsistent results. Disease pressure: the observed
However, there is a robust body of work that clearly argues that invasive species have unique level of disease impact experienced
by a host population, integrating host
roles in disease transmission and are not interchangeable with native species [20]. This curious
susceptibility, pathogen abundance,
mismatch between the foci of disease ecologists and invasion biologists leaves an important gap
virulence, opportunity for
in our understanding of how non-native species affect biodiversity–disease relationships. transmission, and suitable
environmental conditions; frequently
used in an agricultural context.
We thus consider a critical question: how do biological introductions affect parasite richness and
Diversity: a measure of biological
prevalence at the host community (local) level and how does this alter our understanding of
diversity that includes some aspect of
biodiversity–disease relationships? To address this question, we disentangle the multiple and species richness often integrated with
measures of species evenness (e.g.,
often opposing mechanisms by which introduced host species change parasite ecology
Shannon diversity index).
and transmission using both plant and animal systems. We focus first on local-scale mecha-
Enemy escape hypothesis:
nisms by which introduced species impact infectious disease in a community via changes in host
introduced species might be more
diversity, parasite diversity, and host composition. We then consider the impacts of global-scale successful as invaders when they
leave some of their parasites and
patterns of biodiversity homogenization via the proliferation of introduced species, particularly
pathogens in their native range, either
human commensals.
by chance (founder effect) or
because the parasites are unable to
Effects of Introduced Species on Parasitism via Changes in Alpha Diversity thrive in the new environment.
Evenness: a measure of biological
Effects of Changes in Host Species Richness
diversity related to the similarity of
Species richness of the host community is a common way to measure alpha diversity and is
relative abundances across species
– fi
the dominant metric used in analysis of biodiversity disease relationships [11]. Thus, a key rst in an area.
question is whether biological introductions generally lead to net declines (via loss of native Frequency-dependent
transmission: transmission is a
species) or net increases (via the addition of introduced species) in host richness.
function of the proportion of infected
and susceptible host individuals in a
Unfortunately, the answer to this question is complex. Whether and how introduced species population, often through a fixed
number of contacts among hosts or
change the species richness of communities is hotly debated and probably varies across scales
between hosts and vectors;
and systems [21–24]. One complicating issue is that many studies focus on how introduced
examples include sexually transmitted
species affect the richness of native species only, whereas biodiversity–disease relationships
and vector-borne diseases.
concern the richness of the whole community. In addition, the abundance and richness of Habitat filtering: species are able to
establish in some habitats but not
introduced species in a community can be a response to (rather than a cause of) anthropogenic
others based on traits associated
disturbance [25], which independently affects host richness and disease patterns. Thus, one
with physiological tolerance essential
must first quantify the extent to which introduced species drive changes in host richness on a
for them to thrive under particular
landscape and then isolate the effects of such changes on infectious disease from those of other conditions. When those traits are
phylogenetically conserved, habitat
aspects of disturbance.
filtering restricts the phylogenetic
diversity of species in that habitat.
It is also unclear how species introductions affect the total abundance of individuals. Commu- Human commensal: species that
nities are often assumed to be filled to capacity, a zero-sum assumption [26], such that the live in close association with humans
and benefit from the close
introduction of new species causes parallel declines in the abundance of native species. In some
relationship (also called synanthropic
systems, however, introduced species might enter empty niches or communities might be
species).
unsaturated [27]. The extent to which the zero-sum assumption holds is an important knowl- Introduced species: a species
edge gap because it might alter the effects of introduced species on both host richness and introduced through human activities
outside its native distributional range
abundance, which then affect parasitism.
(also sometimes called non-native,
alien, non-indigenous, or exotic).
Assuming the zero-sum assumption is true, we expect that when introduced species increase Invasive species: introduced
host richness, they should, on average, cause declines in both the absolute and relative species that spread aggressively and
cause or have the potential to cause
abundance of native species. As transmission is typically higher within than between host
harm to the environment or to human
species – due to behavioral, physiological, or immunological similarities – such a decrease in the society.
relative abundance of each species should dilute the transmission of generalist parasites with Naturalization hypothesis: species
that are less closely related to native
frequency-dependent transmission (Figure 1) [19]. This effect should be stronger among
species are more likely to succeed as
more specialized parasites [28]. Reductions in absolute abundance per species should also
Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 43
dampen transmission (and prevalence) for specialist parasites with density-dependent invaders by avoiding competition with
transmission. closely related and thus generally
ecologically similar species.
Outbreak: parasite prevalence in
By the same mechanisms and with the zero-sum assumption, introduced species could amplify excess of what would normally be
fi
parasite transmission when they are aggressive invasive species that decrease local host expected in a de ned community or
geographical area.
richness [21,24,29,30]. In this case, the average abundance of parasites per host (including
Parasite: a consumer that lives and
invasive hosts) will increase because a greater proportion of species will be conspecifics and
feeds on a host resulting in harm to
fi
most transmission will occur within species (Figure 1, ampli cation). the host; this is used broadly to
include both microparasites
(pathogens) and macroparasites such
However, if the zero-sum assumption is violated and introduced species are simply added to the
as helminths and arthropods.
local assemblage, the total abundance of native species would not change but their relative
Pathogen: a microparasite (e.g.,
abundance would decrease, causing dilution via encounter reduction. However, because of the virus, bacterium, fungus, protozoan)
that causes disease in its host.
greater total absolute abundance of potential hosts on the landscape, this could potentially be
Preadaptation hypothesis: species
offset by increases in the prevalence of generalist parasites.
that are more closely related to a
dominant native species will be better
Importantly, all of these mechanisms assume that introduced species are, on average, not adapted to a novel environment,
facilitating successful invasion.
systematically different in competence, density, or phylogenetic uniqueness compared with an
Prevalence: a measure of infection
average native species. As described below, these assumptions are unlikely to be generally true.
level that refers to the percentage of
individuals in a population or group
Effects of Changes in Host Species Evenness that are infected with a parasite.
Richness: a measure of biodiversity
In addition to species richness, many metrics of diversity also include a measure of evenness.
focused on a count of the number of
Parasite transmission should decrease strongly with increasing host evenness when host
species, either of hosts or parasites.
abundance is held constant because as evenness increases, both frequency-dependent Spillback: spread of a native
and density-dependent intraspecific transmission should decline, if within-species transmission pathogen in an introduced host and
movement back to the original host,
is greater than between-species transmission [19].
resulting in higher prevalence than
expected in the absence of the new
Most biological communities are very uneven, comprising a few common species and a long tail host.
of rare species. Introduced species tend to make communities even more uneven [29]. For Spillover: spread of a parasite from
one host species to another, atypical
example, in a global network of 64 grasslands, introduced species were four times more likely
host. In the context of this review,
than native species to dominate community cover, whereas natives were 50% more likely than
spillover usually involves transmission
introduced species to be among the rarest species [31]. An invaded grassland is thus typically of novel parasites from introduced
hosts to native hosts, although it can
dominated by one or a few species with native species sharing the remaining space [21].
also involve transmission of native
Assuming that intrinsic susceptibility and competence for any single pathogen are similar
parasites from native hosts to
between native and introduced species (but see below), invaded communities should have
introduced hosts.
a larger proportion of species at low abundances, reducing the probability of outbreaks of Susceptible host: species that can
be infected by a parasite but are not
parasites among native hosts. However, the high density of invasive hosts might conversely
necessarily suitable for parasite
facilitate outbreaks within those hosts (Figure 1, outbreak). Furthermore, when parasites are
reproduction and transmission (in
shared between native and invasive species, a high abundance of the invader can also amplify
contrast to a competent host).
these parasites within native hosts (see spillback, below). Tolerance: ability to minimize the
disease impacts of parasite infection,
often through some form of host
If a greater proportion of the invaded community is dominated by a few (invasive) species, native
defense. Tolerant hosts can be
species will be so rare as to reduce or eliminate transmission of their specialist parasites. competent hosts, able to support
Moreover, since specialist parasites will become extinct before their hosts, declines (or extinc- parasite reproduction and
transmission.
tions) of native host species should drive loss of native parasite richness (Figure 1, coextinc-
Virulence: the severity of disease
tion). A range of studies has demonstrated declines in parasite richness or prevalence when
symptoms (including death) among
native hosts decline in abundance [32,33]. Coextinctions are likely to occur more frequently infected hosts resulting in decreased
fi
with parasites that exhibit complex life cycles that can be disrupted by the loss of any of several host tness.
Zero-sum assumption: an
hosts [34].
assumption that the total number of
individuals in a community is constant
Increasing Parasite Richness via Introduced Hosts and at some carrying capacity.
Introduced host species might also increase community-level parasite richness by introducing
novel parasites (Figure 1, spillover) [35]. Some introduced parasites have caused devastating
44 Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 Efects of invasives
Nave Invasive
G parasites parasites L l o o
c Parasite b
a Invader Nave community Examples a
l Rich Prev
l s
s
c Diversity c
a Decrease in nave plant a
l Diluon e l
e diversity caused increase + in pathogen prevalence Host rich. Invasive grass has – increased parasite Amplificaon infecon in invasive monoculture areas Invasive zebra mussels in zebra mussel dominated Outbreak areas have high trematode loads Host evenness Invasive snails replace Coexncon nave snail and associated Exnct parasite fauna Invasive squirrel Spillover replaces nave squirrel via parapoxvirus Parasite rich. Pathogenic fungi found on one of two nave grasses, Enemy escape but not on exoc grass Composion Invasive gecko Spillback/Amp increases parasite load + in nave gecko Host comp. Introduced trout leads to – decreased parasite load Diluon in nave fauna Introduced plants Habitat filtering decreased phylogenec + diversity, likely Phylogenec div. increasing disease risk – Phylogenecally rare Rare advantage species escaped disease pressure in grassland Homogenizaon Widespread invasive bullfrog transmits invasive fungus to other amphibians Commensal rat facilitated Commensals spread of plague to
North American rodents
Figure 1. Introduced Host Species Can Impact Communities in Multiple Ways, with Diverse Effects on Community-Scale Pathogen Richness and
Prevalence. Arrows denote direction of effects occurring at a local scale (host richness, host evenness, pathogen richness, host competence, phylogenetic diversity) or
(Figure legend continued on the bottom of the next page.)
Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 45
Box 2. Disease-Mediated Invasions
In disease-mediated invasions, parasites play a critical role in facilitating invasions, with the introduced hosts tending to
increase both parasite richness and, often, prevalence while also causing declines in the abundance or diversity of native
hosts. Both parasite spillover and spillback can give the non-native species an advantage over ecologically similar native
species that might otherwise exclude them through competition [35]. For example, North American grey squirrels
(Sciurus carolinensis) have replaced the native red squirrel (Sciurus vulgaris) in the UK aided by mortality caused by the
squirrel parapoxvirus introduced to red squirrels with the arrival of the grey squirrel (see Figure 1 in main text) [84].
Similarly, a key factor in the success of the North American crayfish as an invader in Europe is due to the decimation of
European crayfish populations through transmission of Aphanomyces astacii, the causative agent of crayfish plague,
which is likely to have been co-introduced [74]. In California, the success of European annual grasses is facilitated by
barley yellow dwarf virus, which has stronger effects on Californian native bunchgrasses than on the European species
[85]. Interestingly, disease-mediated invasions inadvertently played a part in shaping aspects of human history. While the
spread of Europeans across the Americas and their decimation of Native American populations in the 15th and 16th
century can be attributed to many factors, diseases such as smallpox that the Europeans introduced to the local
populations certainly played an important role in catalyzing the invasion [35,86].
The mechanism behind disease-mediated invasion can sometimes be less direct, as for the invasive harlequin ladybird
(Harmonia axyridis). Harlequin ladybirds carry parasitic microsporidia that have no effect on them but are lethal to the
native ladybird Coccinella septempunctata [87]. Competing ladybird species commonly feed on the others’ eggs;
however, this competitive behavior ultimately leads to the death of any C. septempunctata ladybirds that feed on H.
axyridis eggs and as a result become infected with the microsporidia [87].
In faunal invasions, parasites have been shown to facilitate invasion not only by the disease they cause but also by the
differential immune investment in introduced species. Because introduced species in a novel environment are less likely
to encounter host-specific pathogens (e.g., through enemy escape), successful invaders might avoid the unnecessary
costs of heightened immune system inflammatory responses to fight disease and instead use less costly antibody
defenses [65]. By saving resources on immune investment, and instead investing in growth and reproduction, the
differential costs of parasitism and disease for introduced versus native species could facilitate the invasion of the
introduced species [65].
impacts through parasite spillover [36]. For instance, Batrachochytrium dendrobatidis, the
pathogen associated with global amphibian declines and extinctions, is partly attributed to
the introduction of non-native amphibians via international trade [37]. Humans have also served
as invasive hosts for pathogens that have then had substantial spillover impacts. For example,
yellow fever, schistosomiasis, and malaria are likely to have been introduced with the African
slave trade to the New World [38–40], where they flourished in new vectors and intermediate
hosts, spreading across the Americas and causing severe disease (and local extinction) in
primate hosts [41]. Parasite spillover might even facilitate invasion via disease-mediated
invasions (Box 2) and is sufficiently common that infectious disease is considered to be the
main driver of negative impacts of invasion on native hosts in 25% of the IUCN list of the worst
invasive species [20].
Notably, not all spillover events have negative consequences for native hosts, even when they
increase both the richness and the prevalence of parasites in a system. If the parasite has a
stronger impact on its original introduced host than on susceptible native species, the parasite
can slow an invasion or diminish its impacts. For instance, the spillover of the fungal seed
pathogen black fingers of death (Pyrenophora semeniperda) from invasive cheatgrass (Bromus
tectorum) facilitates the persistence of native bunchgrasses in the face of widespread cheat-
grass invasion because mortality caused by infection is stronger in the seeds of the invader than
in the native bunchgrasses [42,43]. However, in a review of the virulence of parasites of
a global scale (homogenization and commensals). For each mechanism we provide an example. Where we were unable to document an existing example that
demonstrates the complete mechanism using invasive species (including a measure of consequences at the community scale), we have instead provided an example
showing the mechanism with native species (e.g., dilution via changes in host richness) or the mechanism without documenting community-level consequences in a
single study (e.g., habitat filtering); in these cases we have highlighted the examples box in pink. More details on the examples, including references and appropriate image
attributions, are available in the supplemental information online.
46 Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1
introduced animal species that spill over to native hosts, 85% had higher virulence in native hosts
than in introduced hosts [44]. There appears to be no similar review of the relative virulence of
pathogens in native versus introduced plants. Differences in immune systems between plants
and animals might drive systematic variation between these taxa.
Although the introduction of parasites through new hosts can be devastating to particular native
species, the effects of introduced species on total parasite prevalence at the community level
might often be offset by the fact that, in general, introduced hosts bring with them only a subset
of the total complement of their native parasites (Figure 1, enemy escape) [45]. This observation
has led to the ‘enemy escape hypothesis’ whereby introduced species leave some of their
natural enemies behind; these species can enjoy demographic benefits that enable them to be
more successful as invaders (called ‘enemy release’) [46] (but see [47,48]). With introduced
species accompanied by only a subset of their home-range parasites [45,49], invaded com-
munities might have a lower average number of parasites per host. However, this pattern will be
weakened if introduced hosts acquire native parasites and if parasites from the native range are
subsequently introduced to the newly invaded range [50].
Effects of Invasions on Disease via Changes in Local Species Composition
Biodiversity change involves not only changes in the numbers of species but also functional
differences in the traits and behaviors of introduced compared with native species, and the
potential loss of particular functional roles when native species decline following invasion. We
must consider how introductions that change community composition can also change the
distribution of traits that affect the ability of parasites to reproduce or cause disease.
Effects of Invasion on Abundance of Competent Hosts
In the study of loss of native biodiversity, biologists agree that species loss is nonrandom,
disproportionately affecting species with certain traits, functions, and abundances (e.g., large
body size, low natural abundance) [1]. Aggressive invasive species also often have characteristic
trait and functional profiles. Whether the traits of the invasive species make them more
competent (‘amplifying’) or less competent (‘diluting’) hosts for a given parasite will shape
the direction and magnitude of systematic changes in disease dynamics.
Both classical theory and empirical data suggest that invasive species have a suite of life-history
traits that favor fast growth and reproduction [51–53]. Evolution in the introduced range might
also favor these traits [54]. These are the same traits that are proposed to select for reduced
investment in immune function (animals) or enemy defense (plants), thus resulting in higher
competence for many parasites [55]. Empirical evidence for these hypothesized linkages
includes studies (within native communities) of correlations between fast life histories and higher
parasite competence in both plant and animal systems [56,57], although not all studies find such
links [58]. Common garden experiments have also demonstrated lower defenses against
pathogens in introduced plant populations [59]. A few studies also suggested that introduced
hosts can be preferred hosts for some parasites [60,61]. However, the generality of the
hypothesis that invasive species are more competent for more parasites has not been exten-
sively studied.
When introduced species are especially competent hosts for a local parasite, they will tend to
amplify disease when native parasites increase following infection of the invasive species and
then ‘spillback’ onto native hosts (Figure 1, spillback). If such high competence persists for
multiple parasites, overall community-level parasite prevalence would increase [43,62]. Such
spillback could then cause feedbacks that facilitate invasion. For instance, the introduced
variegated leafhopper has successfully invaded California vineyards because it serves as an
important reservoir host for a native parasitoid (Anagrus epos) that then preferentially attacks the
Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 47
native leafhopper (Erythroneura elegantula) [63]. The generality and average effects of amplifi-
cation and spillback are unclear, however, as relatively few examples exist.
An alternative mechanism has also received support: host resistance to infection might facilitate
colonization success if invasive species have systematically different immune functions [64] or
invest more in immune defense [65], allowing them to better resist native parasites and making
them more likely to be diluting rather than amplifying hosts (Figure 1, dilution). Examples in which
introduced species serve as diluting hosts for native animal parasites have been found in many
ecosystems, including helminth infections in fish [66], Bartonella in rodents [67], and trematodes
in snails [33] and mussels [68]. Introduced hosts should on average also be poor hosts for local
parasites that are host specialists, creating a dilution effect for these parasites. The importance of
introduced species as diluting versus amplifying hosts might thus hinge on the relative impor-
tance of specialist and generalist parasites [48], which is an area ripe for future research and
closely tied to the phylogenetic structure of communities.
Effects from Invasion-Driven Changes in Community Phylogenetic Diversity
The phylogenetic distance between introduced species and the native community might be
especially important for understanding links between invasions and disease dynamics [69].
Many parasites are shared more easily among closely related host species [69–71] and the
relative competence of host species for a given parasite typically shows a phylogenetic signal
[69,72]. As a result, parasite spillover or spillback is more likely between closely related species.
Introduced species that are more closely related to resident species are more likely to encounter
novel parasites to which they are susceptible (facilitating spillback) and parasites hitchhiking on
the introduced species are more likely to spread to novel hosts in the new habitat (facilitating
spillover) [69].
Whether an introduced species is ‘more of the same’ or a phylogenetic novelty is likely to
drive much of the relationship between biological invasion and disease (Box 3). Where
the dominant force in community assembly is habitat filtering based on conserved traits
(Box 3), introduced species are more likely to be close relatives of residents, to be
amplifying hosts for resident parasites, and to share their parasites with native species
(Figure 1, habitat filtering). By contrast, where introduced species are phylogenetically distant
from the resident community they are more likely to dilute parasite transmission (Figure 1,
rare advantage). Consistent with this prediction, the total amount of disease experienced by
each host species in a grassland community was best explained by the abundance of all
species in the community weighted by their phylogenetic distance to that host; novel hosts
introduced to the community suffered less disease when they were more phylogenetically
isolated [69]. Phylogenetically distant species might also be more likely to occupy a new
niche, violating zero-sum assumptions and increasing prevalence via increases in host
density.
Effects of Large-Scale Changes in Diversity and Composition
Global Homogenization
One consistent effect of biological introductions is the global homogenization of flora and fauna
[73], which has massive effects on beta diversity. In a world where species occur not within
biogeographic realms but within global climatic envelopes [74] and where transcontinental
movement of hosts and parasites is relatively common [3], local transmission dynamics might
influence global patterns of disease. Thus, a widely introduced host has many opportunities to
acquire novel parasites from different local pools. Such novel infections are more likely to spread
globally, both because the host species often has numerous dispersal mechanisms (hence its
global distribution) and because conspecifics are present to amplify the parasite when it arrives,
thus increasing both parasite richness and prevalence (Figure 1, homogenization). For example,
48 Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1
Box 3. Darwin's Phylogenetic Conundrum
In On the Origin of Species, Darwin [88] noted contrasting predictions for the role of phylogenetic relatedness in biological
invasions [89]. He first posited a ‘preadaptation hypothesis’, which suggests that successful invaders should be
closely related to dominant species in the landscape because they would more likely have those traits that allow them to
succeed in the new habitat. However, Darwin then favored a ‘naturalization hypothesis’ proposing that species that
are more phylogenetically and phenotypically dissimilar to native species are most likely to become invasive because they
would be more successful competitors among a newly encountered ensemble of species [88]. Essentially, the
naturalization hypothesis proposes that interspecific competition or apparent competition through shared enemies is
the primary element in successful invasion. Consistent with Darwin's naturalization hypothesis, some studies have found
that more phylogenetically distinct species are more likely to be invasive [90,91].
By contrast, the preadaptation hypothesis puts greater weight on what is now called ‘habitat filtering’, where traits
associated with physiological tolerance are the essential keys for successful establishment of introduced species. When
such traits are phylogenetically conserved, this restricts the phylogenetic range of species that thrive in a particular
habitat. Consistent with this, empirical studies have shown that many introduced species have close relatives within the
resident species pool [92,93]. A recent analysis found that introduced plants tend to be as phylogenetically close to native
species in the same region as natives are among themselves [53].
In reality, of course, these hypotheses are not mutually exclusive. For instance, when the plant Ambrosia artemisiifolia was
introduced into 369 different experimental plant communities it was more likely to invade those communities to which it
was phylogenetically closest [94] (Figure I). However, after invading it grew largest where resident species were
phylogenetically distant.
Overall, the phylogenetically conserved traits that shape how a species interacts with its abiotic environment, neighbors,
and parasites are likely to have large and somewhat predictable effects on the performance and impacts of an introduced
species in a new habitat. However, not all ecologically important traits are phylogenetically conserved [95]. Also, because
both habitat filtering and biotic interactions with the local community are always in play in shaping ecological commu-
nities, their relative strength, scale, and timing will influence how invasions change disease dynamics.
Phylogene cally Phylogene cally close invader distant invader
Invader seed introduc on
Preadapta on efect Seed establishment
Efects on disease
Parasite
Figure I. Diagram Showing the Potential Effects of Phylogenetic Relatedness of Introduced Species at
Different Stages of Introduction and Effects on Community Composition [94] and Community-Level
Parasite Abundance for Native Parasites.
Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 49
the invasive bullfrog Lithobates catesbeina is a competent host for the pathogen B. dendro- Outstanding Questions
batidis and is thought to play an important role in transmission of this pathogen [75,76]. Do introduced species tend to cause
decreases or increases in host species
richness and diversity or community
Global Dominance of Human Domestics, Commensals, and Their Parasites
phylogenetic diversity?
Modern humans spread globally within about 70 000 years [77]. Associated with this unprec-
edented global spread of the human mammal was a range of newly domesticated plants and Are there predictable differences in dis-
animals as well as many human commensal species that benefited from pilfering stores of food (i. ease outcomes between invasive and
noninvasive introduced species that
e., rodents) or from the protection of living in human homes and fields (various bats, birds,
are important for disease–diversity
insects, cultivated plants, and weeds). These changes had major impacts on beta diversity on a relationships?
broad scale and often altered local communities of organisms.
What are the properties of invaded
ecosystems that determine the relative
Today, many of the world's most pervasive invasive species are human commensals and
importance of mechanisms related to
agricultural weeds. Probably due to the close and repeated physical interactions between
infectious disease establishment and
humans and these particular introduced hosts, they tend to be very effective at sharing parasites spread?
with humans and their crops (Figure 1, commensals) [78]. Thus, movement of commensal or
domesticated organisms often results in movement of infectious disease to new human Do introduced hosts tend to show high
or low competence for native parasites
populations and to wildlife, as seen in the case of Yersinia pestis (the causative agent of plague)
compared with native hosts? Does this
and its spread around the world by commensal rodents and in the catastrophic effects of
vary between plants and animals?
rinderpest introduced with domestic cattle [79,80]. Similarly, the potato late-blight pathogen
(Phytophthora infestans) evolved in association with species of Solanaceae in North America
What are the net effects of introduced
[81]. After switching onto the potato host, it was transported to Europe on infected potatoes, species at the community scale on both
parasite richness and prevalence?
where it caused the Irish potato blight that contributed to widespread famine and emigration in
1845–1849. More recently, genotypes of P. infestans are thought to have been transported
How do the importance of each mech-
across the world with European seed potatoes and have infected wild plants through a process
anism and the net effects vary between
of hybridization [81]. Thus, commensals are likely to drive a strong positive relationship between frequency- and density-dependent
biological introductions and the number and prevalence of parasites of human relevance – the pathogens?
diseases of crops, livestock, commensals, and humans.
How do differences in pathogen
defense systems in plants and immune
Notably, parasites of human commensals also spill over to affect native species. For instance,
systems in animals drive systematic var-
commensal rats are likely to have facilitated the spread of plague into North American rodents iation between these two groups of
such as prairie dogs, causing dramatic changes in the composition and ecology of many organisms?
ecosystems [82]. Similarly, native rats (Rattus macleari) on Christmas Island in the Indian Ocean
are likely to have been driven to extinction by the introduction of the black rat (Rattus rattus) and
its trypanosome (Trypanosoma lewisi) during the development of phosphate mining on the
island in 1899 [83].
Concluding Remarks: Future Directions
Introduced species are an increasingly dominant part of many landscapes, particularly in
disturbed landscapes where humans and their domesticates are most abundant. Changes
in species composition, driven in large part by introduced species, are likely to be a much more
pervasive signature of anthropogenic effects on biodiversity than any of the commonly used
diversity metrics in disease–biodiversity discussions [5,16].
Because introduced species have the potential to abruptly inject great novelty into a local
disease landscape, they should not be ignored. At the same time, they are not necessarily
equivalent to native species because they are likely to have unique features that change
biodiversity–disease relationships. They can also drive unpredictable effects on community
structure and disease (Box 4).
Unfortunately, based on our current understanding, predicting how introduced species will affect
disease remains fraught because so many factors that matter to pathogen spread have not been
compared systematically between native and introduced species (Figure 1) (see Outstanding
50 Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1
Box 4. Complex and Unpredictable Effects of Invaders on Disease
Invasive species can affect disease dynamics through direct effects on communities (see Figure 1 in main text), but
invaders can also influence disease via indirect and unpredictable pathways [96–98]. In other words, the effects of
introducing a species can cause cascades – as documented for species loss – with unpredictable effects on hosts and
infectious agents. For example, the invasive Amur honeysuckle (Lonicera maacki) indirectly increases human risk of
exposure to a tick-borne disease [96]. Amur honeysuckle has invaded forests throughout North America, many of which
are inhabited by white-tailed deer (Odocoileus virginianus), the primary host for the lone star tick (Amblyomma
americanum) that vectors the causative agents of human ehrlichiosis. Deer prefer shelter and food provided in
honeysuckle-invaded habitat, which leads to greatly elevated densities of white-tailed deer and infected tick nymphs
in honeysuckle-invaded areas [96].
Conversely, the invasive Japanese stiltgrass (Microstegium vimineum) reduces vector tick populations by diminishing
habitat quality [97]. Following experimental introduction of stiltgrass, invaded plots had lower densities of the lone star tick
and the American dog tick (Dermancentor variabilis, vector of Rocky Mountain spotted fever) [97]. This negative effect is
likely to occur because Japanese stiltgrass reduces the relative humidity and increases the temperature of the soil
microhabitat, both of which reduce tick survival [97].
The indirect effects of invaders on disease can reach beyond modifying host availability and habitat quality. For instance,
the gallflies Urophora affinis and Urophora quadrifaciata were intentionally introduced to North America as biological
control agents for the invasive weed spotted knapweed (Centaurea maculosa). However, the intended weed control was
ineffective and Urophora gallflies now infest densely abundant spotted knapweed populations [98,99]. The knapweed
invasion has extended into the native range of deer mice (Peromyscus maniculatus), which are the primary reservoir for
Sin Nombre virus (SNV), the causative agent of hantavirus pulmonary syndrome in humans [100]. In invaded grasslands,
deer mice enjoy a food subsidy of abundant gallfly larvae, which leads to increased survival of deer mice and doubling of
their population size (Figure I) [98]. Mice in invaded areas are also more likely to become infected by SNV, suggesting that
Urophora food subsidies can increase the prevalence of SNV, potentially via density effects (Figure I) [98].
Key: Urophora larvae High C. maculosa 18 Deer mice Low C. maculosa 300 SNV-infected deer mice 16 250 14 2 e m
c 12 200 5 . n 0 a
d /
n 10 e u a b v r a
150 a l e
c i
8 a r m o
r h e p e 6 100 o r D U 4 50 2
0 0
1999 2000 2001 2002 2003 Year
2
Figure I. Mean ( SE) for Density of Introduced Urophora Larvae per 0 5 m , Abundance of Deer Mice
Æ
(Peromyscus maniculatus), and Abundance of Sin Nombre Virus (SNV)-Infected Deer Mice at Sites in
Western Montana with High and Low Invasive Centaurea maculosa Density. Note that the Urophora produced
in one year subsidize deer mice in the next year, because food is consumed over winter. Reproduced using data from
[98].
Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 51
Questions). These factors might also vary among introduced species. For instance, introduced
species might lose coevolved pathogens through enemy escape but might also carry other
pathogens that spill over to native hosts. Introduced species might serve as amplifying hosts if
their life history traits make them competent for multiple resident pathogens or as diluting hosts if
they tend to be phylogenetically distinct from their new neighbors.
Part of the difficulty in resolving the net effects of introduced species on disease stems from
the remaining uncertainty in the net impact of introduced species on different aspects of
biodiversity. However, the largest outstanding challenge is the lack of integration of the
multiple effects of introduced species into a comprehensive predictive framework. This gap
results in part because research programs on disease ecology and on biological introduc-
tions have developed independently, with different underlying goals. Studies of the impacts
of introduced species on diseases of plants and animals typically focus on the conservation
impacts of invasions and document a single mechanism for specific parasites in a subset of
native host species. Likewise, most studies on diversity–disease relationships focus exclu-
sively on native diversity, also usually with a single mechanism and single parasite, typically
with an interest in understanding how conservation of native diversity might protect eco-
systems or human health.
Research on biological introductions and disease ecology needs to be integrated. Especially
with introduced species, multiple mechanisms will operate simultaneously to influence biodi-
versity–disease relationships. We must design studies that explicitly examine the relative
importance of various mechanisms of introduced species on host–parasite interactions, for
multiple types of parasites and across biotic and anthropogenic contexts. Developing this
integrated framework will be critical in any effort to envision and shape how biodiversity change
– and associated conservation actions – will alter future disease landscapes.
Acknowledgments
The authors thank the National Science Foundation (DEB-0842059, DEB-1136626, and DEB-1457371), Shannon Lynch,
Jessica Gee, Collin McCabe, and Maggie Klope. Andy D.M. Dobson, Paul Craze, and two anonymous reviewers provided
helpful comments on the manuscript.
Supplemental Information
Supplemental information associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.tree. 2016.09.008.
References
1. Dirzo, R. et al. (2014) Defaunation in the Anthropocene. Science 10. Wood, C.L. et al. (2014) Does biodiversity protect humans
345, 401–406 against infectious disease? Ecology 95, 817–832
2. Roy, B.A. et al. (2014) Increasing forest loss worldwide from 11. Civitello, D.J. et al. (2015) Biodiversity inhibits parasites: broad
invasive pests requires new trade regulations. Front. Ecol. Envi- evidence for the dilution effect. Proc. Natl Acad. Sci. U.S.A. 112,
ron. 12, 457–465 8667–8671
3. Hulme, P.E. (2009) Trade, transport and trouble: managing inva- 12. Keesing, F. and Ostfeld, R.S. (2015) Is biodiversity good for your
sive species pathways in an era of globalization. J. Appl. Ecol. 46, health? Science 349, 235–236
10–18
13. Johnson, P.T.J. et al. (2015) Frontiers in research on biodiversity
4. Wardle, D.A. et al. (2011) Terrestrial ecosystem responses to and disease. Ecol. Lett. 18, 1119–1133
species gains and losses. Science 332, 1273–1277
14. Sobral, F.L. et al. (2016) Introductions do not compensate for
5. Dornelas, M. et al. (2014) Assemblage time series reveal biodi- functional and phylogenetic losses following extinctions in insular
versity change but not systematic loss. Science 344, 296–299 bird assemblages. Ecol. Lett. 19, 1091–1100
6. Hooper, D.U. et al. (2012) A global synthesis reveals biodiversity 15. Sax, D.F. and Gaines, S.D. (2008) Species invasions and extinc-
loss as a major driver of ecosystem change. Nature 486, 105–108 tion: the future of native biodiversity on islands. Proc. Natl. Acad.
Sci. U.S.A. 105, 11490–11497
7. Taylor, L.H. et al. (2001) Risk factors for human disease emer-
gence. Philos. Trans. R. Soc. Lond. B. Biol. Sci. 356, 983–989 16. Vellend, M. et al. (2013) Global meta-analysis reveals no net
change in local-scale plant biodiversity over time. Proc. Natl.
8. Strange, R.N. and Scott, P.R. (2005) Plant disease: a threat to
Acad. Sci. U.S.A. 110, 19456–19459
global food security. Annu. Rev. Phytopathol. 43, 83–116
17. Winter, M. et al. (2009) Plant extinctions and introductions lead to
9. Smith, K.F. et al. (2006) Evidence for the role of infectious disease
phylogenetic and taxonomic homogenization of the European
in species extinction and endangerment. Conserv. Biol. 20,
fl
1349–1357 ora. Proc. Natl. Acad. Sci. U.S.A. 106, 21721–21725
52 Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1
18. Johnson, P.T.J. and Thieltges, D.W. (2010) Diversity, decoys and 43. Beckstead, J. et al. (2010) Cheatgrass facilitates spillover of a seed
the dilution effect: how ecological communities affect disease bank pathogen onto native grass species. J. Ecol. 98, 168–177
risk. J. Exp. Biol. 213, 961–970
44. Lymbery, A.J. et al. (2014) Co-invaders: the effects of alien
19. Ostfeld, R.S. and Keesing, F. (2012) Effects of host diversity on parasites on native hosts. Int. J. Parasitol. Parasites Wildl. 3,
infectious disease. Annu. Rev. Ecol. Evol. Syst. 43, 157–182 171–177
20. Hatcher, M.J. et al. (2012) Disease emergence and invasions. 45. Torchin, M.E. and Mitchell, C.E. (2004) Parasites, pathogens,
Funct. Ecol. 26, 1275–1287 and invasions by plants and animals. Front. Ecol. Environ. 2, 183–
190
21. Powell, K.I. et al. (2013) Invasive plants have scale-dependent
effects on diversity by altering species–area relationships. Sci- 46. Keane, R.M. and Crawley, M.J. (2002) Exotic plant invasions and
ence 339, 316–318 the enemy release hypothesis. Trends Ecol. Evol. 17, 164–170
22. Thomas, C.D. and Palmer, G. (2015) Non-native plants add to 47. Dostál, P. et al. (2013) Enemy damage of exotic plant species is
the British flora without negative consequences for native diver- similar to that of natives and increases with productivity. J. Ecol.
sity. Proc. Natl. Acad. Sci. U.S.A. 112, 4387–4392 101, 388–399
23. Hulme, P.E. et al. (2015) Challenging the view that invasive 48. Parker, I.M. and Gilbert, G.S. (2007) When there is no escape: the
non-native plants are not a significant threat to the floristic effects of natural enemies on native, invasive, and noninvasive
diversity of Great Britain. Proc. Natl. Acad. Sci. U.S.A. 112, plants. Ecology 88, 1210–1224
E2988–E2989
49. Heger, T. and Jeschke, J.M. (2014) The enemy release hypothe-
24. Vilà, M. et al. (2011) Ecological impacts of invasive alien plants: a sis as a hierarchy of hypotheses. Oikos 123, 741–750
meta-analysis of their effects on species, communities and eco-
50. Mitchell, C.E. et al. (2010) Controls on pathogen species richness
systems. Ecol. Lett. 14, 702–708
in plants’ introduced and native ranges: roles of residence time,
25. MacDougall, A.S. et al. (2014) Anthropogenic-based regional- range size and host traits. Ecol. Lett. 13, 1525–1535
scale factors most consistently explain plot-level exotic diversity
51. Van Kleunen, M. et al. (2010) A meta-analysis of trait differences
in grasslands. Glob. Ecol. Biogeogr. 23, 802–810
between invasive and non-invasive plant species. Ecol. Lett. 13,
26. Hubbell, S.P. (2001) The Unified Neutral Theory of Biodiversity 235–245
and Biogeography, Princeton University Press
52. Baker, H. (1965) Characteristics and mode of origin of weeds. In
27. Brym, Z.T. et al. (2011) Plant functional traits suggest novel TheGenetics ofColonizing Species, pp. 147–172, Academic Press
ecological strategy for an invasive shrub in an understorey woody
53. Ordonez, A. (2013) Functional and phylogenetic similarity of alien
plant community. J. Appl. Ecol. 48, 1098–1106
plants to co-occurring natives. Ecology 95, 1191–1202
28. Holt, R.D. et al. (2003) Parasite establishment in host communi-
54. Dlugosch, K.M. and Parker, I.M. (2008) Invading populations of
ties. Ecol. Lett. 6, 837–842
an ornamental shrub show rapid life history evolution despite
29. Hejda, M. et al. (2009) Impact of invasive plants on the species genetic bottlenecks. Ecol. Lett. 11, 701–709
richness, diversity and composition of invaded communities.
55. Han, B.A. et al. (2015) Rodent reservoirs of future zoonotic
J. Ecol. 97, 393–403
diseases. Proc. Natl. Acad. Sci. U.S.A. 112, 7039–7044
30. Powell, K.I. et al. (2011) A synthesis of plant invasion
56. Ostfeld, R.S. et al. (2014) Life history and demographic drivers of
effects on biodiversity across spatial scales. Am. J. Bot.
reservoir competence for three tick-borne zoonotic pathogens.
98, 539–548
PLoS ONE 9, e107387
31. Seabloom, E.W. et al. (2015) Plant species’ origin predicts dom-
57. Cronin, J.P. et al. (2010) Host physiological phenotype explains
inance and response to nutrient enrichment and herbivores in
pathogen reservoir potential. Ecol. Lett. 13, 1221–1232
global grasslands. Nat. Commun. 6, 7710
58. Luis, A.D. et al. (2015) Network analysis of host–virus communi-
32. Colwell, R.K. et al. (2012) Coextinction and persistence of depen-
ties in bats and rodents reveals determinants of cross-species
dent species in a changing world. Annu. Rev. Ecol. Evol. Syst. 43,
transmission. Ecol. Lett. 18, 1153–1162
183–203
59. Wolfe, L.M. et al. (2004) Increased susceptibility to enemies
33. Torchin, M.E. et al. (2005) Differential parasitism of native and
following introduction in the invasive plant Silene latifolia. Ecol.
introduced snails: replacement of a parasite fauna. Biol. Invasions
Lett. 7, 813–820
7, 885–894
60. Frankel, V.M. et al. (2015) Host preference of an introduced
34. Wood, C.L. et al. (2014) Fishing drives declines in fish parasite
“generalist” parasite for a non-native host. Int. J. Parasitol. 45,
diversity and has variable effects on parasite abundance. Ecology 703–709
95, 1929–1946
61. Parker, J.D. and Hay, M.E. (2005) Biotic resistance to plant
35. Strauss, A. et al. (2012) Invading with biological weapons: the
invasions? Native herbivores prefer non-native plants. Ecol. Lett.
importance of disease-mediated invasions. Funct. Ecol. 26,
8, 959–967
1249–1261
62. Joseph, M.B. et al. (2013) Does life history mediate changing
36. Power, A.G. and Mitchell, C.E. (2004) Pathogen spillover in
disease risk when communities disassemble? Ecol. Lett. 16,
disease epidemics. Am. Nat. 164, S79–S89 1405–1412
37. Fisher, M.C. and Garner, T.W.J. (2007) The relationship between
63. Settle, W.H. and Wilson, L.T. (1990) Invasion by the variegated
the emergence of Batrachochytrium dendrobatidis, the interna-
leafhopper and biotic interactions: parasitism, competition, and
tional trade in amphibians and introduced amphibian species.
apparent competition. Ecology 71, 1461–1470
Fungal Biol. Rev. 21, 2–9
64. White, T.A. and Perkins, S.E. (2012) The ecoimmunology of
38. Bryant, J.E. et al. (2007) Out of Africa: a molecular perspective on
invasive species. Funct. Ecol. 26, 1313–1323
the introduction of yellow fever virus into the Americas. PLoS
65. Lee, K.A. and Klasing, K.C. (2004) A role for immunology in
Pathog. 3, e75
invasion biology. Trends Ecol. Evol. 19, 523–529
39. Yalcindag, E. et al. (2012) Multiple independent introductions of
66. Kelly, D.W. et al. (2009) Has the introduction of brown trout
Plasmodium falciparum in South America. Proc. Natl. Acad. Sci.
altered disease patterns in native New Zealand fish? Freshw.
U.S.A. 109, 511–516
Biol. 54, 1805–1818
40. Crellen, T. et al. (2016) Whole genome resequencing of the
67. Telfer, S. et al. (2005) Disruption of a host–parasite system
human parasite Schistosoma mansoni reveals population history
following the introduction of an exotic host species. Parasitology
and effects of selection. Sci. Rep. 6, 20954
130, 661–668
41. Nunn, C.L. and Altizer, S. (2006) Infectious Diseases in Primates:
68. Thieltges, D.W. et al. (2008) Invaders interfere with native para-
Behavior, Ecology and Evolution, Oxford University Press
site–host interactions. Biol. Invasions 11, 1421–1429
42. Mordecai, E.A. (2013) Despite spillover, a shared pathogen
69. Parker, I.M. et al. (2015) Phylogenetic structure and host
promotes native plant persistence in a cheatgrass-invaded
abundance drive disease pressure in communities. Nature
grassland. Ecology 94, 2744–2753
520, 542–544
Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1 53
70. Cooper, N. et al. (2012) Phylogenetic host specificity and under- 86. Riley, J.C. (2010) Smallpox and American Indians revisited.
standing parasite sharing in primates. Ecol. Lett. 15, 1370–1377 J. Hist. Med. Allied Sci. 65, 445–477
71. Gilbert, G.S. and Webb, C.O. (2007) Phylogenetic signal in plant 87. Vilcinskas, A. et al. (2013) Invasive harlequin ladybird carries
pathogen–host range. Proc. Natl. Acad. Sci. U.S.A. 104, 4979– biological weapons against native competitors. Science 340,
4983 862–863
72. Gilbert, G.S. et al. (2015) The impact of plant enemies shows a 88. Darwin, C. (1859) On the Origin of Species by Means of Natural
phylogenetic signal. PLoS ONE 10, e0123758 Selection, or the Preservation of Favoured Races in the Struggle
for Life, John Murray
73. Sax, D.F. and Gaines, S.D. (2003) Species diversity: from global
decreases to local increases. Trends Ecol. Evol. 18, 561–566 89. Jones, E.I. et al. (2013) Revisiting Darwin's conundrum reveals a
twist on the relationship between phylogenetic distance and
74. Capinha, C. et al. (2013) Effects of climate change, invasive
invasibility. Proc. Natl. Acad. Sci. U.S.A. 110, 20627–20632
species, and disease on the distribution of native European
crayfishes. Conserv. Biol. 27, 731–740 90. Strauss, S.Y. et al. (2006) Exotic taxa less related to native
species are more invasive. Proc. Natl. Acad. Sci. U.S.A. 103,
75. Bonizzoni, M. et al. (2013) The invasive mosquito species Aedes
5841–5845
albopictus: current knowledge and future perspectives. Trends
Parasitol. 29, 460–468 91. Van Wilgen, N.J. and Richardson, D.M. (2011) Is phylogenetic
relatedness to native species important for the establishment of
76. Garner, T.W.J. et al. (2006) The emerging amphibian pathogen
reptiles introduced to California and Florida? Divers. Distrib. 17,
Batrachochytrium dendrobatidis globally infects introduced pop-
172–181
ulations of the North American bullfrog, Rana catesbeiana. Biol.
Lett. 2, 455–459 92. Tingley, R. et al. (2011) Establishment success of introduced
amphibians increases in the presence of congeneric species.
77. Oppenheimer, S. (2012) Out-of-Africa, the peopling of continents
Am. Nat. 177, 382–388
and islands: tracing uniparental gene trees across the map.
Philos. Trans. R. Soc. Lond. B Biol. Sci. 367, 770–784 93. Maitner, B.S. et al. (2012) Patterns of bird invasion are consistent
with environmental filtering. Ecography 35, 614–623
78. McFarlane, R. et al. (2012) Synanthropy of wild mammals as a
determinant of emerging infectious diseases in the Asian–Aus- 94. Li, S. et al. (2015) Contrasting effects of phylogenetic relatedness
tralasian region. Ecohealth 9, 24–35 on plant invader success in experimental grassland communities.
J. Appl. Ecol. 52, 89–99
79. Schmid, B.V. et al. (2015) Climate-driven introduction of the
Black Death and successive plague reintroductions into Europe. 95. Losos, J.B. (2008) Phylogenetic niche conservatism, phyloge-
Proc. Natl. Acad. Sci. U.S.A. 112, 3020–3025 netic signal and the relationship between phylogenetic related-
ness and ecological similarity among species. Ecol. Lett. 11,
80. Dobson, A. et al. (2011) Rinderpest. In Encyclopedia of Biological
995–1003
Invasions (Simberloff, D. and Rejmánek, M., eds), pp. 597–604,
University of California Press 96. Allan, B.F. et al. (2010) Invasive honeysuckle eradication reduces
tick-borne disease risk by altering host dynamics. Proc. Natl.
81. Goss, E.M. et al. (2014) The Irish potato famine pathogen Phy-
Acad. Sci. U.S.A. 107, 18523–18527
tophthora infestans originated in central Mexico rather than the
Andes. Proc. Natl. Acad. Sci. U.S.A. 111, 8791–8796 97. Civitello, D.J. et al. (2008) Exotic grass invasion reduces survival
of Amblyomma americanum and Dermacentor variabilis ticks.
82. Biggins, D.E. and Kosoy, M.Y. (2001) Influences of introduced
J. Med. Entomol. 45, 867–872
plague on North American mammals: implications from ecology
of plague in Asia. J. Mammal. 82, 906–916 98. Pearson, D.E. and Callaway, R.M. (2006) Biological control
agents elevate hantavirus by subsidizing deer mouse popula-
83. Wyatt, K.B. et al. (2008) Historical mammal extinction on Christ-
tions. Ecol. Lett. 9, 443–450
mas Island (Indian Ocean) correlates with introduced infectious
disease. PLoS ONE 3, e3602 99. Myers, J.H. and Harris, P. (1980) Distribution of Urophora galls in
flower heads of diffuse and spotted knapweed in British Colum-
84. Collins, L.M. et al. (2014) Squirrelpox virus: assessing preva-
bia. J. Appl. Ecol. 17, 359–367
lence, transmission and environmental degradation. PLoS ONE
9, e89521 100. Childs, J.E. et al. (1994) Serologic and genetic identification of
Peromyscus maniculatus as the primary rodent reservoir for a
85. Borer, E.T. et al. (2007) Pathogen-induced reversal of native
new hantavirus in the southwestern United States. J. Infect. Dis.
dominance in a grassland community. Proc. Natl. Acad. Sci.
169, 1271–1280
U.S.A. 104, 5473–5478
54 Trends in Ecology & Evolution, January 2017, Vol. 32, No. 1