Emerging infectious disease and the loss of
biodiversity in a Neotropical amphibian community
Karen R. Lips*†, Forrest Brem*, Roberto Brenes*, John D. Reeve*, Ross A. Alford‡, Jamie Voyles§, Cynthia Carey§,
Lauren Livo§, Allan P. Pessier¶, and James P. Collins?
*Department of Zoology, Southern Illinois University, Carbondale, IL 62901-6501;‡School of Tropical Biology, James Cook University, Townsville,
Queensland 4811, Australia;§Department of Integrative Physiology, University of Colorado, Boulder, CO 80309-0354;¶Division of Pathology,
Conservation and Research for Endangered Species, Zoological Society of San Diego, San Diego, CA 92112-0551; and?School of Life Sciences,
Arizona State University, Tempe, AZ 85287-4501
Edited by David B. Wake, University of California, Berkeley, CA, and approved December 26, 2005 (received for review August 9, 2005)
Pathogens rarely cause extinctions of host species, and there are
few examples of a pathogen changing species richness and diver-
sity of an ecological community by causing local extinctions across
a wide range of species. We report the link between the rapid
appearance of a pathogenic chytrid fungus Batrachochytrium den-
drobatidis in an amphibian community at El Cope ´, Panama, and
subsequent mass mortality and loss of amphibian biodiversity
across eight families of frogs and salamanders. We describe an
outbreak of chytridiomycosis in Panama and argue that this infec-
tious disease has played an important role in amphibian popula-
tion declines. The high virulence and large number of potential
hosts of this emerging infectious disease threaten global amphib-
extinction ? fungus ? tropics ? chytridiomycosis ? Panama
infectious diseases are two of eight grand challenges in environ-
mental sciences (1). Emerging infectious diseases may reduce
biodiversity (2) and may account for at least some species
extinctions (3), although, traditionally, infectious disease has not
been considered a cause of extinction, because, in many cases,
the ability of pathogens to be transmitted between hosts (most
specifically R0) is expected to reduce as hosts become rare (4).
Recent modeling of pathogens with multiple hosts, small host-
population sizes, and biotic or abiotic reservoirs, however, has
shown that extinctions of rare species caused by disease may be
more common than realized (4, 5). We provide evidence show-
ing how chytridiomycosis, an emerging infectious disease, is
strongly correlated with decreased population sizes and species
richness of a vertebrate community.
In recent decades, at least 43% of amphibian species have
declined, 32.5% are globally threatened, 34 are extinct, and an
additional 88 are possibly extinct (6). Rapidly declining species
are commonly found in upland Neotropical or Paleotropical
riparian habitats, often in protected areas. These declines were
characterized (6) as ‘‘enigmatic’’ for lack of obvious cause (e.g.,
deforestation or introduced predators), although chytridiomy-
cosis was suspected in many of these cases.
The best documented declines of amphibians are from Central
America, where species richness is high, and endemism is
concentrated in upland areas (7). Whenever mass die-offs were
observed, population declines were rapid (4–6 months), ?50%
of species were extirpated, remaining species persisted at ?20%
of normal abundance, and recovery time exceeded 15 y postde-
cline (8). In most cases where dead and dying frogs were
observed and subsequent population declines occurred, Batra-
chochytrium dendrobatidis, a pathogenic fungus known to kill
amphibians, was found to be the cause of death for almost all
individuals (8–9). This fungus causes chytridiomycosis, an
emerging infectious disease of amphibians (3, 9, 10) associated
with declines in at least 43 amphibian species in seven Latin
nderstanding the causes and consequences of diminishing
American countries and 93 species worldwide (www.jcu.edu.au?
school?phtm?PHTM?frogs?chyglob.htm), but conclusive evi-
dence of the status of the fungus at sites before die-offs and
declines has been lacking. Several studies (refs. 8 and 11; and K.
R. L., unpublished data) suggested that, for over a decade, the
species richness of amphibian communities was collapsing in a
series of habitats along a north-to-south transect in Central
America (Fig. 1). Our goal was to take advantage of this pattern
and document the prevalence of B. dendrobatidis at a site we
believed should have lacked the pathogen, because it was ahead
of the postulated epidemic wave. Should the pathogen continue
to appear at sites to the southeast, we could then describe the full
pattern of change in pathogen and disease prevalence during an
epidemic and the subsequent change in amphibian diversity.
We present evidence that this emerging infectious disease of
amphibians was absent, at very low prevalence, or present
saprobically in the environment before it abruptly increased in
prevalence in many amphibian species and was followed by
widespread mortality and local population extirpations.
Chytridiomycosis was not detected at our study site during the
initial 4 y of sampling until we found the first infected frog on
September 23, 2004. The first dead amphibian positive for B.
dendrobatidis was found October 4, 2004 (Fig. 2). Subsequent
mortality was high (1–19 dead frogs found per day) until mid-
January, 2005, when overall amphibian abundance was much
reduced (Fig. 3). We found 346 dead anurans and 5 dead
salamanders between October 4, 2004 and February 15, 2005; 340
animals on the four riparian transects and 6 (5 frogs and 1
is published as supporting information on the PNAS web site).
There were no dead caecilians, but these largely subterranean
species are difficult to sample. The dead frogs included 38 species
[57% of the amphibian species from this site (K.R.L., unpublished
data)] from all seven families at the site (Table 1). All but 3 of the
(prevalence, 0.98; 95% confidence interval ? 0.964–0.995); those
3 were too decomposed to confidently diagnose infection, whereas
the skin of the other 315 was heavily infected with B. dendrobatidis.
Another 9 species were positive for B. dendrobatidis from PCR
assays of swabs from live frogs for a total of 47 (70%) infected
species of the 67 known from the site (Table 1; and see Supporting
Text). Six of seven samples from substrates associated with dead
frogs were positive for B. dendrobatidis, as was one of nine hap-
hazardly chosen stream boulders.
Conflict of interest statement: No conflicts declared.
This paper was submitted directly (Track II) to the PNAS office.
See Commentary on page 3011.
†To whom correspondence should be addressed. E-mail: firstname.lastname@example.org.
© 2006 by The National Academy of Sciences of the USA
www.pnas.org?cgi?doi?10.1073?pnas.0506889103 PNAS ?
February 28, 2006 ?
vol. 103 ?
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None of the 1,566 individuals of 59 species sampled before
September, 2004 was infected with B. dendrobatidis (Table 1).
Large-scale sampling in October, 2004 showed that prevalence
was already ?10% in 21 of 27 species sampled that month, and,
We analyzed our data in three ways to ensure that our results
truly demonstrate that the pathogen increased dramatically in
prevalence between the pre- and post-September, 2004 samples.
The first analysis is highly conservative. We assumed that species
could differ in ‘‘normal’’ prevalence of B. dendrobatidis and
calculated one-tailed binomial tests for seven species (see Sup-
porting Text). Within each species, pooling all samples collected
before September, 2004 yielded sample sizes large enough (?58
individuals) to provide the statistical power to reject the null
hypothesis that the true prevalence of B. dendrobatidis was ?5%.
For each of these species, we rejected this hypothesis; if B.
dendrobatidis were present in these species before September,
2004, it occurred at very low prevalence. In the second conser-
vative approach, we compared the upper 95% binomial confi-
dence limits for prevalence before September, 2004 with the
lower 95% binomial confidence limits for prevalence after
August, 2004 for all species that had samples available in both
periods. For 17 species (see Supporting Text) these confidence
limits do not overlap, indicating that prevalence in these species
increased significantly after September, 2004. In our final con-
servative analysis, we examined pooled data before September,
2004 for only 22 species that were sampled both before and after
September, 2004 and for which the lower 95% binomial confi-
dence limit for the prevalence of B. dendrobatidis, calculated for
all samples of each species taken after September, 2004, was
?10% (see Supporting Text). These species form a set in which
all species are susceptible to infection by B. dendrobatidis, and
each had a relatively high prevalence during the period of the
outbreak. Pooling across these species to estimate the maximum
prevalence across species before the outbreak, therefore, ap-
pears to be justified. For these species, all 1,217 individuals
of B. dendrobatidis; the upper 95% binomial confidence limit for
prevalence in this aggregated sample is 0.0030, or 0.30%.
In September, 2004, 1 mo before we first detected chytridio-
mycosis, amphibian density and species richness declined
abruptly in both diurnal and nocturnal riparian amphibian
communities; this decline followed six consecutive years of high
and generally increasing population abundances (Figs. 3 and 4).
In the same period, density and species richness of terrestrial
transects showed little evidence of a decline (Figs. 3 and 4).
a similar pattern of decline, although amphibian density and
species richness were consistently lower in diurnal surveys.
Epidermal hyperplasia and hyperkeratosis associated with
10 wild frogs found dead and examined histologically. Most
stores. Chytridiomycosis lesions were diffuse to multifocally
extensive, especially on skin sections from the ventral body.
These lesions were significantly more severe than the relatively
minimal focal lesions described in infected, but apparently
healthy, bullfrogs (12) and were of a level consistent with those
associated with mortality in natural and experimental infections
(13). Molecular analyses of tissue collected from 38 individuals
of 11 species were negative for ranavirus (A. Picco, personal
communication). These findings, with the lack of evidence for
other significant underlying or concurrent disease, suggest that
death was likely caused by chytridiomycosis. We fulfilled Koch’s
postulates for chytridiomycosis by reisolating B. dendrobatidis
from Colosthethus panamensis that had previously been exposed
to isolate JEL 408 (see Supporting Text). Four animals that were
PCR-positive for B. dendrobatidis were examined histologically.
Three were found to have B. dendrobatidis infections, and one
was too decomposed to accurately interpret histologic findings.
In one of the positive cases, the extent of lesions was consistent
with a lethal infection typical of wild animals. Tissues of nine
individuals analyzed by PCR for ranavirus proved to be negative
(A. Picco, personal communication).
Our results demonstrate that the prevalence of B. dendrobatidis
increased from zero to high prevalence very rapidly at our site,
suggesting that B. dendrobatidis invaded the region, causing an
epizootic. We suggest that chytridiomycosis was associated with
mass mortality and the subsequent decline of amphibian popu-
lations at this site. The high prevalence of B. dendrobatidis in
Fig. 1.Map of Central America, with sites of published population declines; lines represent date and location of reported declines (36).
www.pnas.org?cgi?doi?10.1073?pnas.0506889103Lips et al.
dead animals, no detection of other diseases, and fulfillment of
Koch’s postulates supports high disease-induced mortality. Our
sampling did not detect the ascending phase of the epidemic
curve, which probably occurred during August or September.
The timing of this outbreak is consistent with the hypothesis that,
first in Costa Rica and then in Panama, chytridiomycosis is
moving southeastward, thus allowing us to predict its entry into
communities in central Panama. Other survey and monitoring
efforts to the east of El Cope ´ showed no amphibian population
declines (ref. 14 and K.R.L., unpublished data), and our samples
of 120 amphibians from a site 100 km east of El Cope ´ were
negative for chytridiomycosis in January, 2005.
of potential reservoir taxa that could promote long-term per-
sistence of chytridiomycosis that could drive rare or less resistant
species to extinction (4, 5). Likewise, similar sampling of species
at Santa Fe ´, a postdecline site to the west of El Cope ´, identified
six species (Agalychnis callidryas, Bufo marinus, Hyla micro-
cephala, Leptodactylus pentadactylus, Smilisca phaeota, and Rana
‘‘pipiens sp. E’’) with high population abundance, high preva-
lence of chytridiomycosis, and potential for long-distance dis-
persal; these characteristics make them likely disease reservoirs.
Although many diseases can impact host populations by
causing temporary or permanent (2–5) declines in abundance,
only recently has disease been seen as a possible major cause of
species extinctions. Theoretical work on disease ecology predicts
that, as an epidemic infectious disease reduces the abundance of
its hosts, there is an increase in the relative abundance of
immune individuals, and disease transmission is reduced to zero,
such that the pathogen becomes extinct before the host (4).
Theoretically, conditions such as the presence of biotic or abiotic
reservoirs (15, 16), high transmissibility (15), or impacts on host
fecundity (17) could promote extinctions (5). At El Cope ´, many
species are rare, at least 60% of species are susceptible to
infection, and we have identified potential abiotic (this study)
and biotic reservoirs.
We propose that chytridiomycosis operates similarly to a
typical susceptible, exposed, infectious, recovered (SEIR) model
with delayed density dependence (18). That is, chytridiomycosis
emerges at a site and spreads by a combination of frog-to-frog
and environment-to-frog transmission, as was shown in the
laboratory (19) and in field mesocosms (20). As prevalence
increases within the amphibian community, diseased frogs shed
zoospores into the environment or directly pass them to other
amphibians by contact. In the laboratory, B. dendrobatidis zoo-
spores remain viable for at least 3 mo in sterile sand or bird
feathers (21), and infected frogs and salamanders can carry an
infection 24–220 days before dying, with animals at cooler
temperatures shedding zoospores over a longer period (refs. 22,
23, and V. Miera, unpublished data). If positive environmental
proportion of dead frogs in all captures for both night and day transects
(1998–2005). No dead animals were found on transects until October 4, 2004,
at which time mortality increased until January, 2005, by which time abun-
dance was significantly reduced.
Mortality rates for riparian and terrestrial transects, calculated as the
transects (1998–2005). By using a segmented linear model for the riparian
transects, we found a highly significant difference (?2) in slope (t ? ?24.44,
to only the terrestrial transects (see text for details). Diurnal transects were
significantly lower in density than nocturnal ones (t ? ?13.05, df ? 486, P ?
0.0001 for riparian transects; t ? ?9.11, df ? 212, P ? 0.0001 for terrestrial
Lips et al.
February 28, 2006 ?
vol. 103 ?
no. 9 ?
samples from El Cope ´ were also infectious, then environmental
transmission would be possible. All filtered water samples were
negative, suggesting that, if zoospores are in flowing water, they
occur at very low density. We hypothesize that the presence of
B. dendrobatidis in the environment and the long period of
infectivity of many amphibians promote saturation of the envi-
ronment with zoospores, enabling transmission among species
that partition habitat spatially or temporally and producing the
pattern we observed in which prevalence quickly changed from
very low to very high, followed by widespread mortality.
We used three separate analyses to estimate prevalence,
because simply presenting 95% binomial confidence limits for
samples pooled across species (e.g., Table 1) might be criticized
if species were differentially susceptible to chytridiomycosis;
prevalence estimated for pooled samples would inevitably un-
derstate the true prevalence in more-susceptible species and
overstate it in less-susceptible species. If samples taken before
September, 2004, in which we did not detect chytridiomycosis,
included mostly species with low susceptibilities or species which
are constitutively immune to the disease, using the 95% binomial
confidence limit to estimate the upper bound for true prevalence
in susceptible species would overstate the actual statistical power
available. All analyses demonstrated that prevalence was very
low and increased rapidly to high levels.
Initially, it was hypothesized that this disease was endemic and
had recently emerged (24); but growing evidence indicates that
this fungus has been introduced into other areas (3, 10, 25, 26).
This abrupt change in prevalence of infection and subsequent
die-offs are consistent with the introduction of B. dendrobatidis
from infected sites, followed by increasing prevalence and low-
level mortality and infection of environmental substrates, until
dead frogs occur at detectable levels. This pattern is evident in
our data, with a decline in amphibian abundance ?1 mo before
observing the first dead frog.
The die-off at El Cope ´ occurred during the peak of the rainy
season, a pattern like that observed at three other sites: Las Tablas,
Costa Rica; Fortuna, Panama; and Santa Fe ´, Panama (ref. 8; and
K.R.L., unpublished data). Increased prevalence of chytridiomy-
cosis in wild frogs from Australia during the wet season was also
reported (22). Many montane Neotropical frogs breed during the
obligate aquatic chytrid zoospores and increasing amphibian pop-
for a waterborne, directly transmitted pathogen such as B. dendro-
transmission by physical contact during amplexus and defense of
reproductive resources. Our data show that even completely ter-
restrial species become infected and decline, although at reduced
rates; this slower decline is likely caused by reduced transmission in
the terrestrial environment because of greater variability in tem-
perature and moisture conditions in the forest and canopy. Our
report definitively links the appearance of chytridiomycosis in a
community with subsequent declines attributable to this disease.
We do not have detailed climate data from our study site, but
central Panama had no major climatic anomalies in 2004, and
temperature and rainfall patterns were similar to long-term means
(http:??striweb.si.edu?esp). Our results support a model of am-
phibian declines in which B. dendrobatidis enters and quickly
spreads through communities with no infected individuals. Our
model does not require interactions with climate change to provide
a proximate mechanism to account for local extirpations of mon-
that climate change has increased the susceptibility to chytridio-
mycosis of these many amphibian species or that it has influenced
the tempo or mode of spread, especially if susceptibility to chytrid-
iomycosis is affected strongly by microenvironment (27, 28).
We found no evidence that other agents (e.g., exotic predators,
land-use changes, or commercial overharvesting) were involved in
the massive die-off and population declines at El Cope ´. Toxic
chemicals are an unlikely cause, because none of the 10 dead frogs
from the 2004 die-off examined by a pathologist had lesions typical
of exposure to contaminants that result in detectable histologic
changes. A 1998 survey of air, water, sediment, and frog tissues
found no evidence of chemical contaminants of a level likely to
cause mortality (V. Beasley, personal communication).
There never were industries or large-scale or intensive uses of
agrochemicals in this region, and the study site is at the top of
the watershed on the continental divide. Although air currents
could transport contaminants into the site (29), prevailing winds
do not originate from urban or agricultural areas (http:??
mass mortality event and subsequent decline of amphibian
populations at this site. Bioclimatic modeling (30) suggests that
B. dendrobatidis can survive in many other parts of the globe and
supports the idea that it has likely caused other enigmatic
declines of anurans and salamanders throughout the tropics (6).
It is clear that Neotropical amphibian declines and extinctions
are not an artifact of sampling or natural population fluctuations
Table 1. Sampling effort for B. dendrobatidis by time period
infected Prevalence (95% CI)
Pre die-off totals
23 Sep–2 Oct 2004
Post die-off totals
9 0.10 (0.049–0.189)
Columns indicate technique used for assessing B. dendrobatidis, number of species and individuals examined,
and number found infected. Prevalence includes 95% confidence intervals (CI).
www.pnas.org?cgi?doi?10.1073?pnas.0506889103Lips et al.
(31) but, in at least some cases, are caused by a pathogen, B.
dendrobatidis. We predict the loss of many more amphibian
species from the Neotropics, most immediately from montane
regions directly east of our study site. Chytridiomycosis is an
alarming model system for disease-driven extinction of a high
proportion of the species richness of a regional fauna, even a
significant proportion of an entire class of vertebrates, and,
under these circumstances, it is no longer correct to speak of
global amphibian declines but, more appropriately, of global
El Cope ´ (8) is within Parque Nacional G. D. Omar Torrı ´jos H.
(8° 40? N, 80° 37? 17?? W) in Cocle ´ Province, Panama (Fig. 1).
Mean annual air temperature is between 19°C and 26°C, mean
annual precipitation is 3,500 mm, and mean annual water
temperatures are between 20°C and 22°C (K.R.L., unpublished
data). Meteorological data from Barro Colorado Island (http:??
striweb.si.edu?esp), a lowland site 100 km east of El Cope ´, were
a proxy for comparing weather in 2004 to long-term means.
We surveyed amphibian populations along permanent
transects from 1998 to 2005. Each May to August from 1998 to
2003, we surveyed four stream and three terrestrial transects
twice weekly (nocturnal and diurnal). The four stream transects
were surveyed twice monthly from August, 2003 to September,
2004. We surveyed streams at 1- to 3-d intervals and increased
the frequency of terrestrial surveys after October, 2004, when we
began to find dead frogs.
Between June 6, 1998 and January 14, 2005, we carried out 698
transect surveys (187 km) representing ?1,301 h and 29,645 cap-
tures of amphibians. Transect data were tested for changes in
amphibian abundance and species richness by using nonlinear
regression. We predicted that an abrupt change in amphibian
mortality rates would be detectable by a change in slope of
population density vs. time. The change in slope and the time at
which it occurred were estimated by fitting a segmented linear
model using nonlinear regression techniques (32). The model was
of the form Yij? ?0? ?1(t ? ?) ? ?2sin(t ? ?) ? ?X ? ?, where
Yijis amphibian density at time t, t is days elapsed from first sample
date, X is an indicator variable for nocturnal vs. diurnal observa-
tions, and ? is normally distributed with mean 0 and variance ?2.
The parameter ? is the date when population growth rate changes,
?0is the population density at this time, ?1is the average slope of
the two line segments, and ?2is half the difference in slopes (32),
whereas ? estimates the difference in density between night vs. day.
Amphibian density was estimated by using Y ? ln((C ? 1)?(LP)),
where C is total captures, L is the transect length, and P is the
Institute, Cary, NC) was used to fit the model (33). Replacing the
sine function with one that smoothed the transition between line
segments, in particular the function tanh((t ? ?)??), with ? ? 0.1
(32), improved model convergence. We used a similar model to
estimate species richness, with C equal to the total number of
species encountered. We tested if ?2 were different from zero,
calculated confidence intervals for ?, and tested if ? differed from
zero. The segmented linear model readily converged on a solution
for only riparian data, so we fitted a simple linear model to the
terrestrial data using an indicator variable for nocturnal vs. diurnal
We examined 1,566 individuals sampled between May, 2000
and July, 2004 for the presence of B. dendrobatidis. Skin snips
from 145 voucher specimens were examined histologically, and
100 toe tips collected from 7 species between May to June 2003
PCR (34). Between January and July 2004 we made bimonthly
collections of ?300 skin swabs from amphibians encountered on
transects, and assayed pooled swabs for B. dendrobatidis by using
PCR. Positive pooled samples were subdivided and each spec-
imen analyzed individually. We estimated probabilities of false
negatives for each set of bimonthly samples and for all 1,566
samples collectively (Table 1).
All stream transects were surveyed daily for dead amphibians
beginning in October, 2004. Skin from 249 preserved individuals
collected dead between October and December, 2004 were
microscopically examined for the presence of B. dendrobatidis.
We used PCR to estimate prevalence of B. dendrobatidis in 890
frogs sampled from October to December, 2004 (Table 1).
Between September 6 and December 2, 2004, we collected
frogs, and from nine additional boulders without dead frogs. As
part of a companion study, we quantified suspended solids in the
stream by filtering water through glass fiber filters. To determine
whether zoospores of the fungus were present in the water column,
we analyzed solids on 59 alcohol-preserved filters.
Between September 23 and October 2, 2004, we collected 80
individuals of Colostethus panamensis at El Cope ´. These indi-
viduals were swabbed in the field by investigators wearing fresh
latex gloves, and the swabs were tested by Pisces Molecular for
B. dendrobatidis DNA with PCR. Ten of the moribund and dead
amphibians found between October and December in El Cope ´
were chosen haphazardly for histopathology by a veterinary
transects (1998–2005). By using the segmented linear model for the riparian
model to only the terrestrial transects (see text for details). Diurnal transects
were significantly lower in density than nocturnal transects (t ? ?21.33, df ?
486, P ? 0.0001 for riparian transects; t ? ?6.14, df ? 212, P ? 0.0001 for
Species richness and statistical models for riparian and terrestrial
Lips et al.
February 28, 2006 ?
vol. 103 ?
no. 9 ?
pathologist (A.P.P.) to determine the likely cause of death (see
Supporting Text). We also used PCR (35) to test for ranavirus (A.
Picco, personal communication) in tissues collected from 38
haphazardly chosen dead frogs belonging to 11 species.
To verify that B. dendrobatidis could cause mortality in Pana-
manian amphibians, we collected animals in January, 2005 for
fulfillment of Koch’s postulates. We collected 60 C. panamensis
from Cerro Compana, a site 111 km east of El Cope ´ that was
thought to be free of B. dendrobatidis. These animals all tested
negative for B. dendrobatidis by PCR. Koch’s postulates were
panamensis collected in El Cope ´ (see Supporting Text). Eleven of
the animals dying in these experiments were examined by histopa-
thology for signs of B. dendrobatidis infection; B. dendrobatidis was
reisolated from six individuals, and nine were tested by PCR (35)
for ranavirus (A. Picco, personal communication).
We thank field assistants J. Robertson, S. Rodrı ´guez, H. Ross, M. Ryan,
and L. Witters; R. Iba ´n ˜ez, C. Jaramillo, M. Leone, and O. Arosemena
for assistance in Panama; J. Wood for PCR analyses; A. Picco for
ranavirus analyses; J. Longcore for cultures; J. R. Mendelson, E.
Schauber, and M. Parris for reading previous drafts; park staff at Parque
Nacional Omar Torrijos; M. Doran of the Southern Illinois University
(SIUC) histology lab; and V. Miera, H. McCallum, and M. Parris for
permission to cite unpublished data. Climatic data sets were provided by
the Terrestrial–Environmental Sciences Program of the Smithsonian
Tropical Research Institute. This work was supported by National
Science Foundation Grants IBN 9977073, DEB 0130273, DEB 0213851,
and DEB 0234386 and the Bay and Paul Foundation. Research was
approved by SIUC Institutional Animal Care and Use Committee
(IACUC) (01-010, 01-008), Arizona State University IACUC (03-670R),
University of Colorado IACUC, and Smithsonian Tropical Research
Institute. Research and collecting permits were issued by Autoridad
Nacional del Ambiente.
1. Committee on Grand Challenges in Environmental Sciences (2001) National
Research Council Report: Grand Challenges in Environmental Sciences (Natl.
Acad. Press, Washington, DC).
2. Harvell, C. D., Kim, K., Burkholder, J. M., Colwell, R. R., Epstein, P. R.,
Grimes, D. J., Hofmann, E. E., Lipp, E. K., Osterhaus, A. D., Overstreet, R. M.,
et al. (1999) Science 285, 1505–1510.
3. Daszak, P., Cunningham, A. A. & Hyatt, A. D. (2003) Divers. Distrib. 9,
4. McCallum, H. & Dobson, A. (1995) Trends Ecol. Evol. 10, 190–194.
5. de Castro, F. & Bolker, B. (2005) Ecol. Lett. 8, 117.
6. Stuart, S. N., Chanson, J. S., Cox, N. A., Young, B. E., Rodrigues, A. S.,
Fischman, D. L. & Waller, R. W. (2004) Science 306, 1783–1786.
7. Lips, K. R., Burrowes, P. A., Mendelson, J. R. & Parra-Olea, G. (2005)
Biotropica 37, 222–226.
8. Lips, K. R., Reeve, J. D. & Witters, L. (2003) Conserv. Biol. 17, 1078–1088.
9. Berger, L., Speare, R., Daszak, P., Green, D. E., Cunningham, A. A., Goggin,
C. L., Slocombe, R., Ragan, M. A., Hyatt, A. D., McDonald, K. R., et al. (1998)
Proc. Natl. Acad. Sci. USA 95, 9031–9036.
P. J. & Longcore, J. E. (2003) Molec. Ecol. 112, 395–403.
11. Lips, K. R., Green, D. E. & Papendick, R. (2003) J. Herp. 37, 215–218.
12. Hanselmann, R., Rodriguez, A., Lampo, M., Fajardo-Ramos, L., Aguirre,
A. A., Kilpatrick, A. M., Rodriguez, J. P. & Daszak, P. (2004) Biol. Conserv.
13. Nichols, D. K., Lamirande, E. W., Pessier, A. P. & Longcore, J. E. (2001) J.
Wildl. Dis. 37, 1–11.
14. Condit, R., Robinson, W. D., Iba ´ñez, R., Aguilar, S., Sanjur, A., Martı ´nez, R.,
Stallard, R. F., Garcı ´a, T., Angehr, G. R., Petit, L., et al. (2001) Bioscience 51,
15. Anderson, R. M. & May, R. M. (1992) Infectious Diseases of Humans: Dynamics
and Control. (Oxford Univ. Press, NY).
16. Bowers, R. G. & Begon, M. (1991) J. Theor. Biol. 148, 305–329.
17. McCallum, H. I. (1994) Pac. Conserv. Biol. 1, 107–117.
18. Swinton, J., Woolhouse, M. E., Begon, M. E., Dobson, A. P., Ferroglio, E.,
Grenfell, B. T., Guberti, V., Hails, R. S., Heesterbeek, J. A., Lavazza, A., et al.
(2001) in The Ecology of Wildlife Diseases, eds. Hudson, P. J., Rizzoli, A.,
Grenfell, B. T., Heesterbeek, H. & Dobson, A. P. (Oxford Univ. Press, NY),
19. Davidson, E. W., Parris, M., Collins, J. P., Longcore, J. E., Pessier, A. P. &
Brunner, J. (2003) Copeia 2003, 601–607.
20. Parris, M. J. & Cornelius, T. O. (2004) Ecology. 85, 3385–3395.
21. Johnson, M. L. & Speare, R. (2005) Dis. Aquat. Org. 65, 181–186.
22. Berger, L., Speare, R., Hines, H. B., Marantelli, G., Hyatt, A. D., McDonald,
K. R., Skerratt, L. F., Olsen, V., Clarke, J. M., Gillespie, G., et al. (2004)
Australian Vet. J. 82, 424–439.
23. Berger, L., Hyatt, A. D., Olsen, V., Hengstberger, S. G., Boyle, S., Marantelli,
G., Humphreys, K. & Longcore, J. E. (2002) Dis. Aquat. Org. 48, 213–220.
24. Carey, C. Cohen, N. & Rollins-Smith, L. (1999) Dev. Comp. Immunol. 23,
25. Daszak, P. Cunningham, A. A. & Hyatt, A. D. (2000) Science 287, 443–449.
26. Weldon, C., du Preez, L. H., Hyatt, A. D., Muller, R. & Speare, R. (2004)
Emerg. Infect. Dis. 10, 2100–2105.
27. Pounds, J. A. & Pushendorf, R. (2004) Nature 427, 107–109.
28. Woodhams, D., Alford, R. A. & Marantelli, G. (2003) Dis. Aquat. Org. 55, 65–67.
29. Davidson, C., Shaffer, H. B. & Jennings, M. R. (2002) Conserv. Biol. 16,
30. Ron, S. (2005) Biotropica 37, 209–221.
31. Pechmann, J. H. K. Scott, D. E., Semlitsch, R. D., Caldwell, J. P., Vitt, L. J. &
Gibbons, J. W. (1991) Science 253, 892–895.
32. Seber, G. A. F. & Wild, C. J. (1989) Nonlinear Regression (Wiley, NY) pp.
33. SAS version 9.1. (2004) (SAS Institute Inc. Cary, NC).
34. Annis, S. L., Dastoor, F. P., Ziel, H., Daszak, P. & Longcore, J. (2004) J. Wildl.
Dis. 40, 420–428.
35. Mao, J., Tham, T. N., Gentry, G. A., Aubertin, A. & Chinchar, V. G. (1996)
Virology 216, 431–436.
36. Environmental Systems Research Institute (ESRI) (1993) Digital Chart of the
World (CD-ROM Cartographic Database) (ESRI, Redlands, CA).
www.pnas.org?cgi?doi?10.1073?pnas.0506889103Lips et al.