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Ecology Letters, (2005) 8: 117–126
doi: 10.1111/j.1461-0248.2004.00693.x
REVIEW
Mechanisms of disease-induced extinction
Francisco de Castro* and
Benjamin Bolker
Department of Zoology,
University of Florida, 223
Bartram Hall, Box 118525,
Gainesville, FL 32611-8525, USA
*Correspondence: E-mail:
[email protected]
Abstract
Parasites are important determinants of ecological dynamics. Despite the widespread
perception that parasites (in the broad sense, including microbial pathogens) threaten
species with extinction, the simplest deterministic models of parasite dynamics (i.e. of
specialist parasites with density-dependent transmission) predict that parasites will always
go extinct before their hosts. We review the primary theoretical mechanisms that allow
disease-induced extinction and compare them with the empirical literature on parasitic
threats to populations to assess the importance of different mechanisms in threatening
natural populations. Small pre-epidemic population size and the presence of reservoirs
are the most commonly cited factors for disease-induced extinction in empirical studies.
Keywords
Dynamics, reservoir, frequency-dependent, parasite, spatial, specialist.
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Ecology Letters (2005) 8: 117–126
INTRODUCTION
Parasites and disease are frequently cited as important drivers
of population and community dynamics (Anderson & May
1992; McCallum & Dobson 1995; Levin et al. 1997; Hudson
& Greenman 1998; Norman et al. 1999; Kohler & Hoiland
3 2001; Hudson et al. 2002; Shea & Chesson 2002; MacNeil
et al. 2003). In conservation biology, disease is presented as a
threat to population viability and a contributing factor to
disease extinction (McCallum & Dobson 1995). However,
the simplest disease models – which have formed the
theoretical foundation for the field of disease ecology –
suggest that disease alone cannot drive host populations
extinct (Anderson & May 1992). More specifically, deterministic models of directly transmitted specialist parasites with
density-dependent transmission predict that disease will
always die out when the host population falls below a (nonzero) threshold density, before the host population can go
extinct (Swinton et al. 1998; McCallum et al. 2001); stochastic
models suggest that disease will often go extinct by so-called
Ôfade-outÕ even above this threshold (Bartlett 1960; Black
1966; Keeling & Grenfell 1997). The first part of this paper
reviews the important exceptions to these simple conclusions – the qualitative mechanisms that drive disease-induced
extinction in theoretical models. The second reviews the
existing empirical literature on disease-induced extinction,
and makes a first attempt to assess the relative importance of
the different mechanisms in natural systems.
The literature search was performed in the ISI Web of
Science, selecting any article containing the words [Ôextinct*Õ
AND (ÔdiseaseÕ OR Ôparasit*Õ OR ÔpathogenÕ)] in the title,
abstract or keywords. In total, 336 references were found.
References to the coefficient of extinction of some substance
(typically in the context of human physiology and medicine),
articles that only vaguely mentioned disease as a possible
threat for populations, and those referring to the extinction
of parasites (rather than hosts) were dropped (260 in total).
The remaining articles (76), which are the base of this review,
were classified as either theoretical (33) or empirical (43)
(some quantitative simulation models of specific host–
parasite systems were included in the empirical rather than
the theoretical section). We are aware that the so-called Ôgray
literatureÕ, not covered by the Web of Science, contains many
references on conservation subjects, but we feel confident
that our reference base is representative. We have not
included cases of captive populations, since they are
subjected to conditions very different from wild populations.
THEORETICAL MECHANISMS
The list of adjectives qualifying Ôdisease modelsÕ above –
deterministic, density-dependent, specialist – suggests the
mechanisms that allow diseases to drive their hosts extinct in
models. Disease can drive populations temporarily or
permanently to low numbers or densities, predisposing them
to extinction by demographic stochasticity or Allee effects;
diseases with frequency-dependent or spatial transmission
can remain at high incidence even when populations have
become globally rare; and diseases that can exploit other hosts
(biotic reservoirs) or survive and grow in the environment
(abiotic reservoirs) can remain at high incidence independent
of population crashes in the focal host (Table 1).
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118 F. de Castro and B. Bolker
Table 1 Theoretical mechanisms that could produce diseaseinduced extinction
Mechanism
Predisposing factors
Stochasticity/small
populations
Small pre-epidemic populations
(endangered/endemic species)
Small equilibrium density with
disease present
Transient or periodic low
population minima
Allee effects
Inbreeding effects
Fecundity-reducing parasites
Non-density dependent
Frequency-dependent transmission
transmission/inhomogeneous
Sexually-transmitted diseases
mixing
Vector-borne diseases
Social behaviour
Spatial dynamics/metapopulations
Reservoirs
Reservoirs
Biotic: apparent competitors
Abiotic: amplification in the
environment
Small population sizes/stochasticity
While deterministic models allow host populations to fall to
infinitesimal fractions of an individual without going extinct,
stochastic models more realistically constrain the population
to discrete values. With this change, it is easy to see how
parasites could either drive their hosts directly to extinction,
or drive them to such low densities that they can easily be
driven extinct by demographic fluctuations. Once we
recognize the importance of small population sizes, we can
immediately see a variety of ecological scenarios that will
enhance the risk of disease-induced extinction: for example,
diseases attacking small (endemic or endangered) populations, or the presence of Allee effects (negative growth rate of
populations at low densities, potentially caused by a variety of
social, behavioural, or genetic factors). Lack of genetic
variability, often caused by population bottlenecks, can
reduce the ability of individuals to mount an immune
response, increasing the probability that parasites will spread
in the population. Low genetic variability may thus cause a
positive feedback loop, where pathogens more easily invade
small populations which, in turn, decreases their population
size and genetic variability even more, in a process similar to
Lynch et al.Õs (1995) Ômutational meltdownÕ. We have not
found any direct theoretical analysis of this effect, although
we discuss possible empirical examples below (O’Brien et al.
1985; Sanjayan et al. 1996). Even if host populations naturally
occur at high densities prior to an epidemic, parasites can
transiently or permanently depress population densities.
Permanent depression is most severe in the case of parasites
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that spread efficiently (i.e. that have a high R0) and have
intermediate virulence (as measured by the rate of diseaseinduced host mortality), such that they can kill hosts but not
so quickly that they limit their own reproduction (Anderson
& May 1992). Transient depression of host populations,
either during the troughs of host–parasite cycles or during
the initial host population crash that follows the initial
introduction of a parasite to a host population, is enhanced
by long demographic (population regrowth) time scales
relative to the time scale of disease spread (although this
imbalance would also increase the chance of disease
extinction); long incubation periods; and parasites that affect
host fitness through sterility or decreased fecundity, especially if the pathogen does not cause significant mortality.
Frequency-dependent transmission/inhomogeneous
mixing
The predictions of simple disease models also depend
critically on the assumption of density-dependent transmission (McCallum et al. 2001), which is intimately linked with
the assumption of spatially homogeneous mixing. The
simplest departure from density-dependent transmission is
frequency-dependent transmission, where the spread of
disease is proportional not to the absolute number of
susceptible or infectious hosts but rather to the frequency
of infected individuals in the population; the clearest
example of frequency-dependent transmission is in sexually
transmitted diseases (Thrall et al. 1993; Antonovics et al.
1995), but frequency dependence can also occur because of
territorial or social behaviour (Altizer et al. 2003). Since
frequency-dependent transmission eliminates the host density threshold for parasite establishment, it also allows
parasites to drive their hosts to extinction. Getz & Pickering
(1983) show that frequency-transmitted parasites can coexist
with their hosts if some density-dependent factor (other
than the parasite) controls the host population, especially if
density-dependence is stronger for the healthy class than for
the infected class (Thrall et al. 1993).
A parasite transmitted by a vector also has a form of
frequency-dependent transmission, especially for vectors
that are efficient in finding their target species, maintaining
high transmission rates even at low host populations (Boots
& Sasaki 2003). In classical models of vector-borne disease,
transmission is actually inversely proportional to host density,
since vector populations remain constant and thus vectors
become more concentrated on hosts as host populations
decrease. In principle, this inverse density dependence could
lead to a downward spiral as vector-borne disease increases
in a dwindling host population; however, one has to assume
in this case (unrealistically) that the vector population can
support itself independently of the focal host population,
but still maintain its preference for the focal host.
Mechanisms of disease-induced extinction 119
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Thrall et al. (1993), using the same model for sexually or
vector transmitted diseases, define the conditions that allow
a frequency-transmitted disease to coexist with its host.
Coexistence is impossible if infected individuals do not
reproduce and density-dependent effects are equal for
infected and uninfected hosts. One well-studied class of
vector-transmitted diseases is pollinator-borne fungi, which
also fall under the category of sexually-transmitted diseases,
and typically sterilize their hosts. Ingvarsson & Lundberg
(1993), using a deterministic model of the interaction of the
castrating fungus Ustilago violacea and its host, the perennial
herb Lychnis viscaria, show that the parasite can drive it to
extinction as transmission efficiency approaches 100%.
Alexander & Antonovics (1988) analyse the case of a vectortransmitted pathogen (U. violacea) of the plant Silene alba.
They model transmission as dependent on the frequency of
diseased plants, since the parasite is vector-transmitted
(although they do not model the vector population
explicitly). With a high transmission rate of the pathogen
and low host recruitment, the predicted result is the local
extinction of both host and pathogen populations.
Spatial dynamics
Models with explicit spatial structure in host and parasite
densities also induce a sort of frequency-dependent transmission. Disease transmission in spatial models depends on
local rather than global average population densities, and
local densities of infected individuals often remain high even
after global numbers have crashed, because infection is
clustered within the population. Models of lattice-structured
populations show that a host can be deterministically driven
to extinction by a parasite (Sato et al. 1994; Knudsen &
Schotzko 1999; Haraguchi & Sasaki 2000; Boots & Sasaki
2001). Boots & Sasaki (2001) demonstrate that, in this type
of model, host extinction occurs only if the parasite
significantly reduces host reproduction, and that parasites
that have a relatively small effect on host death rate are more
likely to cause host extinction. Models that consider space as
continuous also show that spatial structure increases the
probability of host extinction (Haydon et al. 2002). Models
that describe space in terms of patches rather than
continuous landscapes also allow for the extinction of host
species if disease increases patch extinction rate above the
colonization rate (of course, these models necessarily
assume that disease can cause or increase the chances of
local extinction within a patch). Metapopulation models of
disease have been of great interest, both because many
species of conservation concern live (naturally or through
human intervention) in patchy landscapes, and because
there has been some concern that conservation measures
designed to increase population persistence by increasing
connectivity among patches may also encourage the
transmission of disease (Hess 1994). However, recent
theoretical studies (Gog et al. 2002; McCallum & Dobson
2002) have suggested that increasing connectivity often
favours species persistence, even if it encourages disease
transmission as a side-effect. The difference between these
recent studies and HessÕ approach is that the recent studies
assume that infections have spilled over from more
abundant alternative hosts (see below, section Specialist
vs. generalist parasites: reservoirs). In that case (frequent for
species of conservation concern), movement between
patches will rarely have a negative impact, even for low
probabilities of external infection.
Spatial models also make qualitatively different predictions about the evolution of host–pathogen interactions, as
compared with models assuming homogeneous mixing.
O’Keefe & Antonovics (2002) analyse the dynamics of
parasites that reduce the fecundity of the host, but not its
survival. According to a deterministic, non-spatial model, if
host fecundity declines with increased pathogen transmission rate, then selection on the pathogen will cause host
fecundity to decline toward zero, causing the extinction of
the host and the pathogen. On the other hand, a spatially
explicit stochastic model with the same assumptions
predicts that virulence evolution does not necessarily lead
to pathogen or host extinction, as it does in a well-mixed
community. Boots & Sasaki (2003) also show that, in
spatially explicit models, although disease-induced extinction may occur, it is unlikely to evolve. Moreover, they show
that there must be a relationship between transmission and
virulence for evolution to endemic coexistence of the host
and the parasite. Haraguchi & Sasaki (2000) show that the
evolutionary stable strategy for transmission and virulence in
a spatial setting may be different than that predicted by a
non-spatial or well-mixed model. The spatial structure leads
to an intermediate transmission rate, even in the absence of
an explicit tradeoff between transmission and virulence. The
possibility of host or parasite extinction or coexistence
depends on the host reproductive rate. Rand et al. (1995)
also predict a substantially lower selection pressure in a
spatial individual-based model, compared with its nonspatial analogue. All of these studies are exploring essentially
the same mechanism: local structure leads to local extinction
and thus to a form of group selection that drives the
evolution of intermediate virulence.
Specialist vs. generalist parasites: reservoirs
It is clear that generalist parasites can overcome host density
thresholds and drive a focal host species to extinction. The
detailed community dynamics of multiple parasites sharing
multiple hosts can be quite complex (Holt & Pickering 1985;
Bowers & Turner 1997; Gatto & De Leo 1998; Bowers &
Hodgkinson 2001; Holt et al. 2003). However, multi-host
parasites can clearly lead to apparent competition, where a
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120 F. de Castro and B. Bolker
host species drives a competitor to extinction by being more
tolerant or encouraging reproduction of a parasite that
harms its competitor (Schmitz & Nudds 1994; Holt et al.
2003). Few theoretical explorations of parasite-mediated
competition have appeared, probably because the mechanism is so simple, but analysing the details of particular
empirical cases can be complicated (e.g. Schmitz & Nudds
1994). In the same general category as these biotic reservoirs
are abiotic reservoirs, where a parasite can also survive and
amplify itself in the environment as a saprophyte (Thrall
et al. 1997). Rosá et al. (2003) show that the presence of an
external reservoir may easily drive the host to extinction if
the level of external infection is high.
Another potential mechanism of extinction is indirect or
trophically mediated extinction; the extinction of one
species by disease, or even a radical change in its population
density, can have cascading effects through an ecological
community (cf. Dunne et al. 2002). For example, in a
reversal of the typical scenario of extinction through
apparent competition, reduced density of one prey species
could lead to predator switching, increased predator
pressure, and hence decreased population density or even
extinction of a second prey species that shared that
predator. Another, even more direct, mechanism is the
reduction of a species that may lead to extinction of a
specialist consumer. Many potential scenarios exist – our
own explorations (F. de Castro, B. Bolker, unpublished
manuscript) have clearly just scratched the surface. In the
next section, we will present a few possible examples from
the existing empirical literature.
EMPIRICAL LITERATURE
We have surveyed the empirical literature on disease
extinction to find cases where species have been threatened
with or driven to extinction by disease. There are very few
cases where disease has been implicated as the direct cause
in the global extinction of a species (Pounds & Crump 1994;
Daszak & Cunningham 1999). In fact, we have found no
examples of the complete extinction of a species in the wild
which can be attributed to a pathogen with certainty.
However, in some cases (Table 2) disease has either
Table 2 Empirical examples of mechanisms of disease-induced extinction
Mechanism
Species
Impact
Reference
Small populations/
stochasticity
Tree snail (Partula turgida)
Thylacine (Thylacinus cynocephalus)
Extinction
Possible extinction
Golden toad (Bufo periglenes)
Black-footed ferret (Mustela nigripes)
Mednyi arctic fox (Alopex lagopus semenovi)
African wild dog (Lycaon pictus)
Boreal toad (Bufo boreas)
Noble crayfish (Astacus astacus)
Spanish Ibex (Capra pyrenaica hispanica)
Probable extinction
Probable extinction
Probable extinction
Population crash
Population crash
Population crash
Population crash
Big-horn sheep (Ovis canadiensis)
Florida torreya (Torreya taxifolia)
Population crash (model)
Population crash and predicted
extinction (model)
Increased susceptibility to disease
Increased susceptibility to disease
Increased susceptibility to disease
Possible population crash (model)
Population crash
Possible population crash (model)
Possible population crash (model)
Population crash
Population crash
Population reduction
Daszak & Cunningham (1999)
Guiler (1961) (in McCallum &
Dobson 1995)
Pounds et al. (1997)
Thorne & Williams (1988)
Goltsman et al. (1996)
Burrows et al. (1994)
Muths et al. (2003)
Taugbol et al. (1993)
Fandos (1991),
Leon-Vizcaino et al. (1999)
Gross et al. (2000)
Schwartz et al. (1995, 2000)
Sanjayan et al. (1996)
Peterson et al. (1998)
O’Brien et al. (1985)
Augustine (1998)
Funayama et al. (2001)
White et al. (2003)
Thrall et al. (2003)
Atkinson et al. (1995)
Atkinson et al. (2000)
Schmitz & Nudds (1994)
Population
Population
Population
Extinction
Population
Rushton et al. (2000)
Haydon et al. (2002)
Richards et al. (1999)
Carlton et al. (1991)
Martinez & Zuberogoitia (2001)
Reduced genetic variability Pocket gopher (Thomomys bottae)
Wolves on Isle Royale (Canis lupis)
Cheetahs (Acinonyx jubatus)
Non-density dependent
Koala/Chlamydia
transmission
Eupatorium makinoi Asteraceae
Inhomogeneous mixing
Rabbit/Rabbit haemorrhagic disease
Common flax (Linum marginale)
Biotic reservoir
Iiwi (Vestiaria coccinea)
Amakihi (Hemignathus virens)
White-tailed deer (Odocoileus virginianus)/
moose (Alces alces)
Red squirrel (Sciurus vulgaris)
Ethiopian wolf (Canis simensis)
Abiotic reservoir
Tussock moth (Orgya antiqua)
Indirect extinction
Eelgrass limpet (Lottia alveus)
Eagle owl (Bubo bubo)
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reduction
reduction (model)
crash
crash
Mechanisms of disease-induced extinction 121
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threatened species or driven local populations extinct
(Thorne & Williams 1988; Schwartz et al. 1995; Goltsman
4 et al. 1996; Richards et al. 1999; Funayama et al. 2001; Muths
et al. 2003).
Small population size
Of the mechanisms discussed in the first section, small
populations are the mostly widely reported driver of diseaseinduced extinction in natural populations, although of
course this could be a bias caused by the higher attention
usually received by endangered species.
Total extinction of a species is cited only in a few cases,
and the role of the pathogen is confirmed in even fewer.
The snail Partula turgida became extinct in 1996 due to a
microsporidian infection (Steinhausia sp., Daszak & Cunningham 1999), although it should be noted that the last
individuals of P. turgida were kept in captivity, which makes
it easier for an epizootic to affect all the remaining
individuals. Another possible case of disease-induced
extinction is the Thylacine (Thylacinus cynocephalus) (Guiler
1961, in McCallum & Dobson 1995). Although hard
evidence is absent, the precipitous nature of the Thylacine’s
decline from 1906 to 1910, its simultaneous occurrence
throughout Tasmania and some anecdotal reports of
animals dying of a Ôdistemper-likeÕ disease, may suggest that
the final extinction (after a long period of over-hunting)
could have been caused by a pathogen. Finally, the golden
toad (Bufo periglenes), an amphibian endemic to Costa Rica
with a very limited distribution, underwent a population
crash in 1987 (Pounds et al. 1997). The species is currently
classified by the IUCN as critically endangered or extinct
(IUCN 1996). Pounds & Crump (1994) suggest the
synergistic interaction of climate variations with a pathogen
as the possible cause of the decline. Lips (1999) suggests
that a fungal infection shared by many other species in
Panama, whose populations are also declining, could have
been the cause of the apparent extinction of the golden
toad.
In other cases, extinction of a species by a disease has
been avoided (or delayed) only by human intervention.
Without that intervention, the characteristics of the disease
made the final extinction probable. One well known
example is the black-footed ferret (Mustela nigripes). The last
known wild colony was extirpated by an epizootic of canine
distemper in 1985 (Thorne & Williams 1988). Sixteen
individuals, captured as an emergency measure during the
epidemic, were devoted to a programme of breeding in
captivity. Without those captive animals, the species would
have probably been driven globally extinct by the disease.
Also, a subspecies of arctic fox, Alopex lagopus semenovi, of
Mednyi Island (Commander Islands, Russia), was reduced
by some 85–90%, to around 90 animals, as a result of mange
introduced by dogs in the 1970s (Goltsman et al. 1996). The
population is currently under treatment with antiparasitic
drugs, but its viability is still uncertain. The extremely high
cub mortality (up to 96%), the social habit of the species
(potentially leading to frequency-dependent transmission)
and the low pre-epizootic population (around 600 individuals) all contributed to make probable the final extinction of
the subspecies.
In many cases, although disease did not cause the total
extinction of the host, it did produce a mortality high
enough to threaten the species or to cause the disappearance
of local populations. Such examples are more frequent, but
it should be noted that any pathogen, almost by definition,
will cause some decrease in the host population density; it is
difficult to draw a clear line separating parasites with the
potential to cause extinction. (Gog et al. 2002 and McCallum
& Dobson 2002 both give lists of pathogens with impacts
on host populations; we have discussed here only those with
the potential to lead to local or global extinction.) Burrows
et al. (1994, 1995) suggest that the local extinction of African
wild dog (Lycaon pictus) in the Serengeti was caused by a viral
disease enhanced by stress due to intervention (vaccination,
collaring, etc.), although Ginsberg et al. (1995) argue that the
population disappearance may have been caused by other
factors. Taugbol et al. (1993) report the disappearance of the
noble crayfish (Astacus astacus) from some lakes and long
stretches of several rivers in Norway due to a plague of a
fungal pathogen (Aphanomyces astaci). The pathogen’s characteristics (species-specific, directly transmitted, short-lived
spores) make complete eradication of the host improbable.
Global declines of several amphibian populations provide
well-documented examples of population extirpation by a
pathogen. The evidence that a fungus (Batrachochytrium
dendrobatidis) is responsible for some of the declines is strong
(see for instance Collins & Storfer 2003 and Daszak et al.
2003). In particular Muths et al. (2003) report the neardisappearance of several populations of boreal toad
(Bufo boreas) in Colorado (USA) apparently due to this same
pathogen, and populations of yellow-legged frogs (Rana
muscosa) are apparently vanishing from the Californian Sierra
Nevada, again apparently from chytrid fungus (C. Briggs,
pers. comm.; Bradford 1991; Fellers et al. 2001). Some
theoretical analyses of specific cases also predict the
extinction of populations or species. For instance Gross
et al. (2000), simulating populations of big-horn sheep (Ovis
canadiensis) with an individual-based model, found a probability of extinction exceeding 20% over a 200 year period
with multiple infections.
As noted above, finite-population effects can also be
important in populations that are initially large, but are
rapidly reduced in density and numbers by disease. For
instance, shortly after 1987, the population of Spanish Ibex
(Capra pyrenaica hispanica) in southern Spain suffered a 95%
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122 F. de Castro and B. Bolker
reduction due to a sarcoptic mange epizootic, from about
9,600 individuals to less than 200 (Fandos 1991), including
the total disappearance of a local group of 35 individuals in
6 months (Leon-Vizcaino et al. 1999).
We have found only a single recent reference to a plant
species threatened with extinction by a disease. The Florida
torreya (Torreya taxifolia) suffered a severe population
reduction during the 1950s, probably as a result of a fungal
disease (Schwartz et al. 1995). The present population is
composed of small, sexually immature trees, not producing
seeds. All the models applied to its population dynamics
predict the extinction of the species (Schwartz et al. 2000).
Two other plant pathogens have also caused dramatic and
widespread declines in their hosts: the chestnut blight
(Endothia parasitica) (Griffin 2000) and the Dutch elm disease
(Ophiostoma ulmi) (Gibbs 1978; Brasier 1987; Hubbes 1999).
Another aspect of the risk associated with small
population size is lack of genetic variability. Extreme
population bottlenecks can reduce the genetic variability
of a species and reduce its effective immune response,
making it more vulnerable to disease. There are several
examples of high mortality attributed to this cause: Cheetahs
(Acinonyx jubatus) and feline infectious peritonitis (O’Brien
et al. 1985); pocket gopher (Thomomys bottae) (Sanjayan et al.
1996); and wolves (Canis lupis) at Isle Royale and canine
parvovirus (Peterson et al. 1998).
Small host populations are the most commonly cited
reason for disease threatening host population viability.
Although only one reference conclusively attributes the
extinction of a whole species (Partula turgida) to disease,
there are two cases (the Thylacine and the golden toad)
where disease is a plausible cause or cofactor, and others
(the black-footed ferret and the Mednyi arctic fox) where
extinction was probably avoided only because of human
intervention. A large variety of other examples cite disease
as a threat, but do not convincingly show that disease alone
will drive extinction. Except in the case of the golden toad,
which was locally abundant before the population crashed,
none of the references cites the disease as the cause of the
initial decline from abundance to scarcity, but rather as a
threat to already small populations.
Frequency-dependent transmission/inhomogeneous
mixing
Frequency-dependent transmission
As discussed above, simple models of host–parasite
dynamics with frequency-dependent transmission (broadly
including STDs and vector-borne diseases) show that a
parasite can easily drive its host to extinction (Getz &
Pickering 1983). Despite this theoretical generalization, we
have found no empirical evidence of a population or species
eliminated by frequency-dependent parasites. Some models
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applied to specific situations suggest that the extinction of
the host is possible (Alexander & Antonovics 1988;
Ingvarsson & Lundberg 1993), but apparently extinction
has never been reported for a real population. A model of
Koala (Phascolarctos cinereus) and the sexually transmitted
disease Chlamydia psittaci shows that extinction of the host is
possible if the transmission of the disease increases and, at
the same time, Koalas become less resistant to the disease or
if KoalasÕ intrinsic population growth rate drops below 0.1
(Augustine 1998). However, stable coexistence of the Koala–
Chlamydia system is predicted for a wide range of population
and disease parameters, and the general consensus seems to
be that Chlamydia has not led to local extinction of koala
populations (Melzer et al. 2000). In general, there is little
empirical evidence that diseases that are transmitted in a
frequency-dependent way are particularly threatening.
Spatial dynamics
Despite the abundant theoretical literature on spatial host–
pathogen interactions, we have found little empirical
evidence on the subject. Thrall et al. (2003) use an
experimental metapopulation approach to study the host–
pathogen system Linum marginale–Melampsora lini, although
they focus in the local extinction and recolonization of the
pathogen, not the host. They show that more isolated
patches tended to exhibit lower levels of disease during
epidemic peaks than patches that were close together. White
et al. (2003), studying rabbit haemorrhagic disease, describe
how the different use of space of surface-dwelling and
warren-based rabbits creates a hierarchical contact and
transmission structure within the rabbit population that
affects the population-level outcome of the infection.
Although the rabbit populations studied did not disappear,
this study is an example of how spatial structure can affect
infection dynamics. The lack of empirical studies on this
subject may be due to the difficulty of managing controlled
experiments of disease dynamics together with spatial
dynamics. Clearly, more efforts are needed in this area to
verify theoretical predictions.
Specialist vs. generalist parasites: reservoirs
The effect of reservoirs in host–parasite dynamics and its
influence on the relationship between hosts is straightforward and theoretically well understood, which is reflected in
the relatively numerous empirical examples of this type of
interaction. The case of avian malaria (Plasmodium relictum)
and its effect on Hawaiian avifauna has been well described.
The Iiwi (Vestiaria coccinea), a declining native species from
Hawaii, is much more susceptible to avian malaria than the
introduced Nutmeg Mannikin (Lonchura punctulata) (Atkinson et al. 1995). (One could argue that the vector transmission of avian malaria also affects its population impact, but
Mechanisms of disease-induced extinction 123
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the simple thought experiment of removing non-native
species – which would harm native species by intensifying
transmission if malaria’s vector transmission is regulating its
impact – suggests that reservoirs are more important.) The
presence of the Mannikins acting as reservoirs for avian
malaria threatens the remaining high elevation populations
of endangered native birds. Another species, the Hawaii
Amakihi (Hemignathus virens) is also restricted in its distribution by its higher susceptibility to malaria (Atkinson et al.
2000). The case of the interaction between the white-tailed
deer (Odocoileus virginianus) and the moose (Alces alces) is also
well known: the white-tailed deer can competitively exclude
moose through its higher resistance to the parasitic
meningeal helminth Parelaphostrongylus tenuis (Anderson
5 1974; Lankester 1987; Schmitz & Nudds 1994). Other
examples include the parapoxvirus-mediated competitive
exclusion of red squirrels by gray squirrels (Rushton et al.
2000); the spillover effects of rabies and canine distemper
from domestic dogs to the threatened Ethiopian wolf
(C. simensis) (Haydon et al. 2002); and the spillover of
rinderpest to wild ruminants from domestic cattle (Rossiter
et al. 1998; Kock et al. 1999).
In some conditions, a reservoir can also allow the
coexistence of host and parasite. Funayama et al. (2001)
describe how a reservoir species (Lonicera japonica) for a virus
affecting the plant Eupatorium makinoi would enable the virus
to persist after the local extinction of the host, allowing their
long-term coexistence, assuming that recolonization by the
host is possible from other populations.
Abiotic reservoirs can also drive extinction, although
theory suggests that either the pathogen must be able to
grow (and not merely persist) in the environment, or that
an interaction with small-population effects is required.
We have found only one reference to the role of an
abiotic reservoir in the disappearance of a population
(Richards et al. 1999). Early-instar larvae of the polyphagous vapourer moth (Orgyia antiqua), parasitized by a
nucleopolyhedrovirus, disperse from tree canopy to the
understorey, with subsequent relocation back to the tree as
late-instar larvae. Virus spores are capable of persisting
outside the host in a viable state for long periods of time
if shielded from sunlight. Widespread epizootics provoked
by larval foraging behaviour seemingly wiped out the
population (meticulous searches of egg masses in the study
area were unsuccessful in the 2 years following the
epizootics).
Trophic and competitive interactions can also allow
diseases to cause the extinction of non-focal species
indirectly, even when their host population persists. If the
disease causes strong enough reduction of the host
population, some species dependent on the host may
become extinct. The eelgrass limpet (Lottia alveus) has
apparently gone extinct as a consequence of the decline of
the eelgrass Zostera marina (Carlton et al. 1991), whose
population suffered a catastrophic decline of in the early
1930s in the North Atlantic Ocean, attributed to the
Ôwasting diseaseÕ of Zostera, which is caused by the slime
mould Labyrinthula zosterae (Muehlstein et al. 1991). Another
example is the severe reduction of the Eagle Owl population
(Bubo bubo) in Alicante (Spain) after an outbreak of rabbit
haemorrhagic disease decimated the population of its main
prey in the area (Martinez & Zuberogoitia 2001). Even if not
causing extinctions, the effects of diseases in keystone
species can be very widespread; for example, the consequences of rabbit population declines due to introduced
myxomatosis (Sumption & Flowerdew 1985).
CONCLUSIONS
We have reviewed the basic ways in which simple
epidemiological models can be extended to allow for
disease-induced extinction of host populations. Despite a
great deal of theoretical exploration, we have found that
some of these mechanisms have few or no confirmed
empirical examples (non-homogeneous mixing, spatial
structure, social behaviour, sexually transmitted diseases).
There is little documentation of the various mechanisms or
predisposing factors other than small population sizes and
reservoirs. Spatial and frequency-dependent mechanisms are
interesting theoretically, but as yet there is little empirical
evidence that they are important in natural systems. In some
sense the paucity of data is not surprising, however, since we
are looking at the intersection of two areas (extinction and
disease ecology) which are both by their nature cryptic and
difficult to study in natural communities.
Another factor to consider is that empirical studies
typically include several of the mechanisms mentioned in
this review: small populations and reservoirs, vector-borne
and spatial dynamics, etc. while theoretical studies tend to
focus in only one mechanism. In studying each empirical case,
careful consideration is needed to include all possible factors
implicated, to obtain a clear understanding of the process.
ACKNOWLEDGEMENTS
NSF Integrated Research Challenges in Environmental
Biology grant IBN 9977063 supported this work. We would
like to thank the comments of the members of our
laboratory and four anonymous referees that greatly
improved the manuscript.
REFERENCES
Alexander, H.M. & Antonovics, J. (1988). Disease spread and
population dynamics of anther-smut infection of Silene alba
caused by the fungus Ustilago violacea. J. Ecol., 76, 91–104.
Ó2004 Blackwell Publishing Ltd/CNRS
124 F. de Castro and B. Bolker
Altizer, S., Nunn, C.L., Thrall, P.H., Gittleman, J.L., Antonovics, J.,
Cunningham, A.A. et al. (2003). Social organization and parasite
risk in mammals: integrating theory and empirical studies. Annu.
Rev. Ecol. Evol. Syst., 34, 517–547.
Anderson, R.C. (1974). The ecological relationships of meningeal
worms and native cervids in North America. J. Wildl. Dis., 8,
304–310.
Anderson, R.M. & May, R.M. (1992). Infectious Diseases of Humans.
Dynamics and Control. Oxford University Press, New York.
Antonovics, J., Iwasa, Y. & Hassell, M.P. (1995). A generalizedmodel of parasitoid, venereal, and vector-based transmission
processes. Am. Nat., 145, 661–675.
Atkinson, C.T., Woods, K.L., Dusek, R.J., Sileo, L.S. & Iko, W.M.
(1995). Wildlife disease and conservation in Hawaii: Pathogenicity
of avian malaria (Plasmodium relictum) in experimentally infected
Iiwi (Vestiaria coccinea). Parasitology, 111 (Suppl. S), S59–S69.
Atkinson, C.T., Dusek, R.J., Woods, K.L. & Iko, W.M. (2000).
Pathogenicity of avian malaria in experimentally-infected Hawaii
Amakihi. J. Wildl. Dis., 36, 197–204.
Augustine, D.J. (1998). Modelling Chlamydia-Koala interactions:
coexistence, population dynamics and conservation implications.
J. Appl. Ecol., 35, 261–272.
Bartlett, M.S. (1960). The critical community size for measles in the
US. J. R. Stat. Soc. A, 123, 37–44.
Black, F.L. (1966). Measles endemicity in insular populations –
Critical community size and its evolutionary implication. J. Theor.
Biol., 11, 207.
Boots, M. & Sasaki, A. (2001). Parasite-driven extinction in spatially explicit host–parasite systems. Am. Nat., 34, 706–713.
Boots, M. & Sasaki, A. (2003). Parasite evolution and extinctions.
Ecol. Lett., 6, 176–182.
Bowers, R.G. & Hodgkinson, D.E. (2001). Community dynamics,
trade-offs, invasion criteria and the evolution of host resistance
to microparasites. J. Theor. Biol., 212, 315–331.
Bowers, R.G. & Turner, J. (1997). Community structure and the
interplay between interspecific infection and competition.
J. Theor. Biol., 187, 95–109.
Bradford, D.F. (1991). Mass mortality and extinction in a highelevation population of Rana muscosa. J. Herpetol., 25, 174–177.
Brasier, C.M. (1987). Recent genetic changes in the Ophiostoma ulmi
populations: the threat to the future of the elm. In: Population of
Plant Pathogens (eds Wolfe, M.S. & Caten, C.E.). Blackwell Scientific Publications, Oxford, pp. 213–226.
Burrows, R., Hofer, H. & East, M.L. (1994). Demography,
extinction and intervention in a small population: the case of the
Serengeti wild dogs. Proc. R. Soc. Lond. B, 256, 281–292.
Burrows, R., Hofer, H. & East, M.L. (1995). Population dynamics,
intervention and survival in African wild dogs (Lycaon pictus).
Proc. R. Soc. Lond. B, 262, 235–245.
Carlton, J.T., Vermeij, G.J., Lindberg, D.R., Carlton, D.A. &
Dudley, E.C. (1991). The 1st historical extinction of a marine
invertebrate in an ocean-basin – the demise of the eelgrass
limpet Lottia alveus. Biol. Bull., 180, 72–80.
Collins, J.P. & Storfer, A. (2003). Global amphibian declines:
sorting the hypotheses. Div. Distrib., 9, 89–98.
Daszak, P. & Cunningham, A.A. (1999). Extinction by infection.
8 Trends Ecol. Evol., 14, 279–279.
Daszak, P., Cunningham, A.A. & Hyatt, A.D. (2003). Infectious
disease and amphibian population declines. Div. Distrib., 9, 141–
150.
Ó2004 Blackwell Publishing Ltd/CNRS
Dunne, J.A., Williams, R.J. & Martinez, N.D. (2002). Network
structure and biodiversity loss in food webs: robustness increases with connectance. Ecol. Lett., 5, 558–567.
Fandos, P. (1991). La cabra montés Capra pyrenaica en el Parque
Natural de las Sierras de Cazorla, Segura y Las Villas. ICONACSIC, Madrid, Spain.
Fellers, G.M., Green, D.E. & Longcore, J.E. (2001). Oral chytridiomycosis in the mountain yellow-legged frog (Rana muscosa).
Copeia, 4, 945–953.
Funayama, S., Terashima, I. & Yahara, T. (2001). Effects of
virus infection and light environment on population
dynamics of Eupatorium makinoi (Asteraceae). Am. J. Botany, 88,
616–622.
Gatto, M. & De Leo, G.A. (1998). Interspecific competition among
macroparasites in a density-dependent host population. J. Math.
Biol., 37, 467–490.
Getz, W.M. & Pickering, J. (1983). Epidemic models: thresholds
and population regulation. Am. Nat., 121, 892–898.
Gibbs, J.N. (1978). Development of the Dutch elm disease in
southern England, 1971–6. Ann. Appl. Biol., 8, 219–228.
Ginsberg, J.R., Alexander, K.A., Creel, S., Kat, P.W., McNutt, J.W.
& Mills, M.G.L. (1995). Handling and survivorship of African
wild dog (Lycaon pictus) in five ecosystems. Conserv. Biol., 9, 665–
674.
Gog, J., Woodroffe, R. & Swinton, J. (2002). Disease in endangered
metapopulations: the importance of alternative hosts. Proc. R.
Soc. Lond. Ser. B, 269, 671–676.
Goltsman, M., Kruchenkova, E.P. & Macdonald, D.W. (1996). The
Mednyi arctic foxes: treating a population imperilled by disease.
Oryx, 30, 251–258.
Griffin, C.J. (2000). Blight control and restoration of the American
chestnut. J. Forestry, 98, 22–27.
Gross, J.E., Singer, F.J. & Moses, M.E. (2000). Effects of disease,
dispersal, and area on bighorn sheep restoration. Restor. Ecol., 8
(4S), 25–37.
Haraguchi, Y. & Sasaki, A. (2000). The evolution of parasite
virulence and transmission rate in a spatially structured population. J. Theor. Biol., 203, 85–96.
Haydon, D.T., Laurenson, M.K. & Sillero-Zubiri, C. (2002). Integrating epidemiology into population viability analysis: managing the risk posed by rabies and canine distemper to the
Ethiopian Wolf. Conserv. Biol., 16, 1372–1385.
Hess, G.R. (1994). Conservation corridors and contagious disease:
a cautionary note. Conserv. Biol., 8, 256–262.
Holt, R.D. & Pickering, J. (1985). Infectious disease and species
coexistence: a model of Lotka–Volterra form. Am. Nat., 126,
196–211.
Holt, R.D., Dobson, A.P., Begon, M., Bowers, R.G. & Schauber,
E.M. (2003). Parasite establishment in host communities. Ecol.
Lett., 6, 837–842.
Hubbes, T. (1999). The American elm and Dutch elm disease.
Forestry Chronicle, 75, 265–273.
Hudson, P.J. & Greenman, J.V. (1998). Competition mediated by
parasites: biological and theoretical progress. Trends Ecol. Evol.,
13, 387–390.
Hudson, P.J., Rizzoli, A., Grenfell, B.T., Heesterbeek, H. &
Dobson, A.P. (2002). The Ecology of Wildlife Diseases. Oxford
University Press, Oxford, pp. 197.
Ingvarsson, P.K. & Lundberg, S. (1993). The effects of a vectorborne disease on the dynamics of natural plant populations: a
1
model for Ustilago violacea infection of Lychnis viscaria. J. Ecol., 81,
263–270.
9 IUCN (1996). Red List Web Site: http://www.redlist.org/.
Keeling, M.J. & Grenfell, B.T. (1997). Disease extinction and
community size: modeling the persistence of measles. Science,
275, 65–67.
Knudsen, G.R. & Schotzko, D.J. (1999). Spatial simulation of
epizootics caused by Beauveria bassiana in Russian wheat aphid
populations. Biol. Control, 16, 318–326.
Kock, R.A., Wambua, J.M., Mwanzia, J., Wamwayi, H., Ndungu,
E.K., Barrett, T. et al. (1999). Rinderpest epidemic in wild
ruminants in Kenya 1993–97. Vet. Rec., 145, 275–283.
Kohler, S.L. & Hoiland, W.K. (2001). Population regulation in an
aquatic insect: the role of disease. Ecology, 82, 2294–2305.
Lankester, M.W. (1987). Pest, parasites and diseases of moose
(Alces alces) in North America. Swedish Wildl. Res., S1, 461–489.
Leon-Vizcaino, L., de Ybanez, M.R.R., Cubero, M.J., Ortiz, J.M.,
Espinosa, J., Perez, L. et al. (1999). Sarcoptic mange in Spanish
ibex from Spain. J. Wildl. Dis., 35, 647–659.
Levin, S.A., Grenfell, B.T., Hastings, A. & Perelson, A.S. (1997).
Mathematical and computational challenges in population biology and ecosystems science. Science, 275, 334–343.
Lips, K.R. (1999). Mass mortality and population declines of anurans
at an upland site in western Panama. Conserv. Biol., 13, 117–125.
Lynch, M., Conery, J. & Burger, R. (1995). Mutation accumulation
and the extinction of small populations. Am. Nat., 146, 489–518.
MacNeil, C., Dick, J.T.A., Hatcher, M.J., Terry, R.S., Smith, J.E. &
Dunn, A.M. (2003). Parasite-mediated predation between native
and invasive amphipods. Proc. R. Soc. Lond. Ser. B, 270, 1309–1314.
Martinez, J.A. & Zuberogoitia, I. (2001). The response of the Eagle
Owl (Bubo bubo) to an outbreak of the rabbit haemorrhagic
disease. J. Ornithol., 142, 204–211.
McCallum, H. & Dobson, A. (1995). Detecting disease and parasite
threats to endangered species and ecosystems. Trends Ecol. Evol.,
10, 190–194.
McCallum, H. & Dobson, A. (2002). Disease, habitat fragmentation and conservation. Proc. R. Soc. Lond. Ser. B, 269, 2041–2049.
McCallum, H., Barlow, N. & Hone, J. (2001). How should pathogen transmission be modelled? Trends Ecol. Evol., 16, 295–
300.
Melzer, A., Carrick, F., Menkhorst, P., Lunney, D., St John, B.
(2000). Overview, critical assessment, and conservation implications of Koala distribution and abundance. Conserv. Biol., 14,
619–628.
Muehlstein, L.K., Porter, D. & Short, F.T. (1991). Labyrinthula
zosterae sp. nov. the causative agent of wasting disease of eelgrass,
Zostera marina. Mycologia, 83, 180–191.
Muths, E., Corn, P.S., Pessier, A.P. & Green, D.E. (2003). Evidence for disease-related amphibian decline in Colorado. Biol.
Conserv., 110, 357–365.
Norman, R., Bowers, R.G., Begon, M. & Hudson, P.J. (1999).
Persistence of tick-horne virus in the presence of multiple host
species: tick reservoirs and parasite mediated competition.
J. Theor. Biol., 200, 111–118.
O’Brien, S.J., Roelke, M.E., Marker, L., Newman, A., Winkler,
C.A., Meltzer, D. et al. (1985). Genetic basis for species
vulnerability in the cheetah. Science, 227, 1428–1434.
O’Keefe, K.J. & Antonovics, J. (2002). Playing by different rules:
the evolution of virulence in sterilizing pathogens. Am. Nat.,
159, 597–605.
Mechanisms of disease-induced extinction 125
Peterson, R.O., Thomas, N.J., Thurber, J.M., Vucetich, J.A. &
Waite, T.A. (1998). Population limitation and the wolves of Isle
Royal. J. Mammal., 79, 828–841.
Pounds, J.A. & Crump, M.L. (1994). Amphibian declines and climate disturbance – the case of the golden toad and the harlequin
frog. Conserv. Biol., 8, 72–85.
Pounds, J.A., Fogden, P.L., Savage, J.M. & Gorman, G.C. (1997).
Test of null models for amphibian declines on a tropical
mountain. Conserv. Biol., 11, 1307–1322.
Rand, D.A., Keeling, M. & Wilson, H.B. (1995). Invasion, stability
and evolution to criticality in spatially extended, artificial host–
pathogen ecologies. Proc. R. Soc. Lond.: Biol. Sci., 259, 55–63.
Richards, A., Cory, J., Speight, M. & Williams, T. (1999). Foraging
in a pathogen reservoir can lead to local host population extinction: a case study of a Lepidoptera-virus interaction. Oecologia, 118, 29–38.
Rosá, R., Pugliese, A., Villani, A. & Rizzoli, A. (2003). Individualbased vs. deterministic models for macroparasites: host cycles
and extinction. Theor. Popul. Biol., 63, 295–307.
Rossiter, P.B., Hussain, M., Raja, R.H., Moghul, W., Khan, Z. &
Broadbent, D.W. (1998). Cattle plague in Shangri-La: observations on a severe outbreak of rinderpest in northern Pakistan
1994–95. Vet. Rec., 143, 39–42.
Rushton, S.P., Lurz, P.W.W., Gurnell, J. & Fuller, R. (2000).
Modelling the spatial dynamics of parapoxvirus disease in red
and grey squirrels: a possible cause of the decline in the red
squirrel in the UK? J. Appl. Ecol., 37, 997–1012.
Sanjayan, M.A., Crooks, K., Zegers, G. & Foran, D. (1996).
Genetic variation and the inmune response in natural populations of pocket gophers. Conserv. Biol., 10, 1519–1527.
Sato, K., Matsuda, H. & Sasaki, A. (1994). Pathogen invasion and
host extinction in lattice structured populations. J. Math. Biol., 32,
251–268.
Schmitz, O.J. & Nudds, T.D. (1994). Parasite-mediated competition in deer and moose: how strong is the effect of meningeal
worm on moose? Ecol. Appl., 4, 91–103.
Schwartz, M.W., Hermann, S.M. & Vogel, C.S. (1995). The catastrophic loss of Torreya taxifolia: assessing environmental induction of disease hypotheses. Ecol. App., 5, 501–516.
Schwartz, M.W., Hermann, S.M. & van Mantgem, P.J. (2000).
Population persistence in Florida torreya: comparing modeled
predictions of a declining coniferous tree. Conserv. Biol., 14,
1023–1033.
Shea, K. & Chesson, P. (2002). Community ecology theory as a
framework for biological invasions. Trends Ecol. Evol., 17, 170–
175.
Sumption, K.J. & Flowerdew, J.R. (1985). The ecological effects of
the decline in rabbits (Oryctolagus cuniculus L.) due to myxomatosis. Mamm. Rev., 15, 151–186.
Swinton, J., Harwood, J., Grenfell, B.T. & Gilligan, C.A. (1998).
Persistence thresholds for phocine distemper virus infection in
harbour seal Phoca vitulina metapopulations. J. Anim. Ecol., 67,
54–68.
Taugbol, T., Skurdal, J. & Hastein, T. (1993). Plague and management strategies in Norway. Biol. Conserv., 63, 75–82.
Thorne, E.T. & Williams, E.S. (1988). Disease and endangered
species: the black-footed ferret as a recent example. Conserv. Biol.,
2, 66–74.
Thrall, P.H., Antonovics, J. & Hall, D.W. (1993). Host and pathogen coexistence in sexually-transmitted and vector-borne
Ó2004 Blackwell Publishing Ltd/CNRS
126 F. de Castro and B. Bolker
diseases characterized by frequency-dependent disease transmission. Am. Nat., 142, 543–552.
Thrall, P.H., Bever, J.D., Mihail, J.D. & Alezander, H.M. (1997).
The population dynamics of annual plants and soil-borne fungal
pathogens. J. Ecol., 85, 313–328.
Thrall, P.H., Godfree, R. & Burdon, J.J. (2003). Influence of spatial
structure on pathogen colonization and extinction: a test using
an experimental metapopulation. Plant Pathol., 52, 350–361.
White, P.C.L., Newton-Cross, G.A., Gray, M., Ashford, R., White,
C. & Saunders, G. (2003). Spatial interactions and habitat use of
Ó2004 Blackwell Publishing Ltd/CNRS
rabbits on pasture and implications for the spread of rabbit
haemorrhagic disease in New South Wales. Wildl. Res., 30, 49–58.
Editor, Masakado Kawata
Manuscript received 2 August 2004
First decision made 30 August 2004
Manuscript accepted 1 October 2004