Content uploaded by Rebecca M. Mccaffery
Author content
All content in this area was uploaded by Rebecca M. Mccaffery on Feb 11, 2014
Content may be subject to copyright.
Decreased winter severity increases viability of
a montane frog population
Rebecca M. McCaffery
a,1
and Bryce A. Maxell
b
a
Wildlife Biology Program, College of Forestry and Conservation, University of Montana, Missoula, MT 59812; and
b
Montana Natural Heritage Program,
University of Montana, Helena, MT 59620
Edited by David B. Wake, University of California, Berkeley, CA, and approved April 6, 2010 (received for review November 9, 2009)
Many proximate causes of global amphibian declines have been
well documented, but the role that climate change has played and
will play in this crisis remains ambiguous for many species. Breeding
phenology and disease outbreaks have been associated with
warming temperatures, but, to date, few studies have evaluated
effects of climate change on individual vital rates and subsequent
population dynamics of amphibians. We evaluated relationships
among local climate variables, annual survival and fecundity, and
population growth rates from a 9-year demographic study of
Columbia spotted frogs (Rana luteiventris) in the Bitterroot Moun-
tains of Montana. We documented an increase in survival and
breeding probability as severity of winter decreased. Therefore,
a warming climate with less severe winters is likely to promote
population viability in this montane frog population. More gener-
ally, amphibians and other ectotherms inhabiting alpine or boreal
habitats at or near their thermal ecological limits may benefit from
the milder winters provided by a warming climate as long as suit-
able habitats remain intact. A more thorough understanding of how
climate change is expected to benefit or harm amphibian popula-
tions at different latitudes and elevations is essential for determin-
ing the best strategies to conserve viable populations and allow
for gene flow and shifts in geographic range.
amphibian
|
climate change
|
demography
|
Rana luteiventris
|
snowpack
Amphibian populations are declining around the globe at an
alarming rate (1–3), and climate change now figures promi-
nently as a potential interactive driver of some of these declines
(4–6). A shift to earlier breeding phenology has been documented
in a number of species (7–11), but this shift is not universal across
species and has not been tied to population-level consequences.
Other work has associated climatic conditions with disease-related
declines in the Neotropics (4, 5), yet no mechanisms have been
definitively linked to these correlations (12, 13). Reading et al. (14)
showed a decrease in adult female body condition and survival in
a population of common toads (Bufo bufo) that corresponded
with an increase in average annual temperatures. However, these
demographic changes were not explicitly linked to changes in
population size over time. Kiesecker et al. (15) showed that dis-
ease, UV-B, and climate change could interact to increase embryo
mortality in western toad (Bufo boreas) populations, yet these
populations have not declined. This increased premetamorphic
mortality may not be sufficient to cause otherwise increasing
populations to decline in the long term. Alternatively, it may be
that conditions in short-term or laboratory-based studies are not
always representative of long-term patterns in natural pop-
ulations. To determine mechanisms of population change, we first
need to know which vital rates are affected by changes in climate,
and then how these changes affect population dynamics.
The effects of climate change on growth and survival are par-
ticularly relevantfor amphibian species in high elevation temperate
ecosystems, where individuals are exposed to extreme and variable
temperatures (e.g., ref. 16). In these environments, amphibians are
active at a wider range of temperatures than closely related species
inhabiting lower elevations (17). In general, juveniles and adults in
high elevation temperate environments must reproduce and ac-
quire resources over much shorter growing seasons than low ele-
vation individuals. Similarly, tadpoles must be able to develop and
metamorphose over a short season or survive long cold winters. In
these environments, longer growing seasons due to climate
warming may allow more time for adults, juveniles, and tadpoles to
grow and acquire resources. Winter severity may decrease, re-
ducing winterkill due to freezing or hypoxia, which can be a major
cause of mortality in these systems (18, 19). However, reduced
precipitation and a warm environment could result in less water in
these high-elevation landscapes, which may negatively impact
amphibian species that rely on ephemeral pools for reproduction
and foraging. Additionally, amphibian populations and speciesthat
are adapted to a colder thermal regime in alpine systems may have
physiological constraints that prevent adaptation to an increase in
temperature (20). With climate change models for the western
United States predicting a reduction in snowpack, as well as in-
creased variability in precipitation levels (21, 22), understanding
the role of these climate variables in the growth and survival of
high-elevation populations of amphibians is critical to predicting
future impacts of climate change on amphibians.
We evaluated relationships between annual vital rate estimates,
estimates of asymptotic population growth rate, and local climate
variables in a high elevation population of a temperate pond-
breeding frog species, the Columbia spotted frog (Rana lutei-
ventris). To evaluate these relationships, we used both mark-re-
capture and demographic analyses. A longstanding principle of
demography is that not all vital rates are equally important for
population dynamics. For example, large changes in annual re-
cruitment, which are commonly documented in amphibians, tend
to have less impact on population growth rates than relatively
small changes in postmetamorphic survival rates (23). We exam-
ined variation in survival, growth, and recruitment in relation to
summer and winter climate variables. We then used matrix models
to calculate asymptotic growth rate (λ) for each set of annual vital
rates (24–26) and related λto these climate variables by using
linear regression. Our results advance our understanding of how
climate change may affect amphibian populations, and have
implications for the conservation of other temperate amphibian
species occupying alpine and boreal habitats.
We conducted our study in the Little Rock Creek Basin, found
in the Selway-Bitterroot Wilderness ≈16 km south of Hamilton,
MT. This drainage is composed of two glacial cirques. Our
analyses were focused on the upper basin (2,200 m), which is
delineated by glacial headwalls (Fig. 1). Although this upper
basin contains multiple breeding and foraging sites, they function
as one population (27). The basin contains requisite habitats for
R. luteiventris overwintering, breeding, and summer foraging, and
Author contribution s: R.M.M. and B.A.M. designed resea rch; R.M.M. and B.A.M. per-
formed research; R.M.M. analyzed data; and R.M.M. and B.A.M. wrote the paper.
The authors declare no conflict of interest.
This article is a PNAS Direct Submission.
1
To whom correspondence should be addressed. E-mail: amphibecs@gmail.com.
This article contains supporting information online at www.pnas.org/lookup/suppl/doi:10.
1073/pnas.0912945107/-/DCSupplemental.
8644–8649
|
PNAS
|
May 11, 2010
|
vol. 107
|
no. 19 www.pnas.org/cgi/doi/10.1073/pnas.0912945107
is fishless (Fig. 1). We have not tested for the presence of the
amphibian chytrid fungus Batrachochytrium dendrobatidis (Bd) in
this population, but Bd is widespread in Montana and has been
detected in amphibians within a few miles of this watershed (28).
However, no mortality events were noted in this watershed, no
local or broad-scale declines have been detected for R. lutei-
ventris in western Montana during the course of this study (28),
and R. luteiventris is known to have skin peptides that are highly
resistant to Bd (29). Therefore, we believe that common culprits
for amphibian declines at high elevation sites (i.e., introduced
fish and disease; refs. 30 and 31) do not complicate our analyses.
We monitored all life stages of R. luteiventris from 2000 to 2008
and related these demographic data to climate data collected at
a nearby weather station. Climate variables included (i) peak
snow-water equivalency (hereafter “SWE”, a common measure
of snowpack), (ii) winter length, (iii) end of winter Julian date
(hereafter “last day of winter”), (iv) summer length, and (v)
growing degree days (see Methods for complete description of
climate variables).
Results
Climate Variables. During the 9 years of our study, SWE ranged
from 67 to 135 cm, winter lengths varied by 41 days, last day of
winter varied by 33 days, summer length ranged from 84 to 120
days, and growing degree days ranged from 431 to 648 days.
There were no trends in any of the climate variables over the
study period. The range of SWE falls within the historical range
of peak SWE at this station (46–171 cm), and average peak SWE
for the length of our study is identical to the historical average
(101 cm). Therefore, the range and magnitude of variability in
climate variables seen in our study captures what has been
recorded historically. Winter variables were correlated with each
other (R
2
>0.62), as were summer variables (R
2
= 0.66), but
winter variables were not correlated with summer variables.
Relationships Between Climate and Vital Rates. Breeding probability
was inversely related to the length of the previous winter (P=
0.002; Fig. 2). There was no relationship between breeding
probability and either of the summer climate variables. There
was no relationship between survival to 1 yr of age and any of the
winter or summer climate variables.
Survival and transition probabilities depended on winter severity.
Capture-recapture data for 4,279 individual frogs captured over
9 years were analyzed by using program MARK (32), with climate
variables included as group covariates in various combinations.
Model selection results for the capture probability models in-
dicated that probability of initial capture (p) and probability of
recapture (c) varied both across years and among life stages.
Therefore, we used this capture probability structure in all sub-
sequent transition and survival models. Capture probabilities
ranged from 4 to 50%, depending on the year and the life stage.
Models with SWE and last day of winter were far better supported
than any other candidate models (AIC weight = 0.999, model
likelihood = 1.0, 93 parameters). In these models, an increase in
SWE was associated with a decrease in survival for juvenile and
adult stage classes (Fig. 3). Juvenile frogs showed the strongest
relationship with variation in SWE (β=−0.042 ±0.005 SE), but
the effect was also significant in adult females (β=−0.037 ±0.011
SE) and adult males (β=−0.051 ±0.012 SE). There was no
relationship between SWE and survival for subadult females
(β=−0.023 ±0.018 SE).
SWE was inversely related to transition from juvenile to sub-
adult female or male (β=−0.029 ±0.007 SE) and transition
from juvenile to adult female (β=−0.081 ±0.024 SE). There
was no relationship with transition from subadult to adult female
(β= 0.015 ±0.015 SE).
Demographic Analysis. We used survival and transition parameter
estimates from a time-varyingcapture-recapture model to calculate
annual vital rates (Table S1) to parameterize female-based, post-
birth pulse Leftkovich matrix models for each year of the study
(Fig. 4). Across these matrices, estimates of λwere higher in years
with lower SWE (P= 0.023; Fig. 5A), tended to be lower in years
with a later last day of winter (P= 0.185; Fig. 5B), and were not
related to winter length (P= 0.611).
Lower Basin
Upper Basin
Fig. 1. The Little Rock Creek drainage, Bitterroot Mountains, Montana. All
analyses were based on data collected in the upper basin. Habitat used by R.
luteiventris includes the following: (i) permanent water bodies with no
emergent vegetation used for summer foraging and overwintering (hatch
marks); (ii) permanent ponds and lakes with emergent vegetation used for
breeding, foraging, and overwintering (solid); and (iii) ephemeral ponds
with emergent vegetation used for breeding and foraging (open).
2001
2002
2003
2004
2005 2006
2007 2008
Fig. 2. Linear regression of female breeding probability with length of
previous winter. Longer winters were associated with lower breeding
probabilities in the following spring (P= 0.002).
0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
65
67
69
71
73
75
77
79
81
83
85
87
89
91
93
95
97
99
101
103
105
107
109
111
113
115
117
119
121
123
125
127
129
131
133
135
137
139
Survival probability
Peak SWE (cm)
Juvenile Survival Subadult Survival Female Survival Male Survival
Fig. 3. Relationship between stage-specific survival rates in R. luteiventris
and SWE for four age/sex classes: juvenile, subadult female, adult male, and
adult female. The lines are the predicted survival curves for different values
of SWE from the top model.
McCaffery and Maxell PNAS
|
May 11, 2010
|
vol. 107
|
no. 19
|
8645
ECOLOGY
Discussion
Parameters that describe winter severity were negatively corre-
lated with survival, transition, and breeding probabilities in this
high elevation R. luteiventris population. These relationships
resulted in higher estimated λin years with earlier ending winters
and lower snowpack. Climate change predictions for the Rocky
Mountains suggest that snowpack will continue to decrease and
snowmelt will occur at an increasingly earlier date (33). Because
effects of decreased winter severity were uniformly positive, our
results indicate that this climate trend will increase population
viability of R. luteiventris in this mountain system. Researchers
have typically assumed that if climate change affects amphibian
species, the outcome will be negative. Rarely do we allow our-
selves to consider that global warming may confer benefits to
some species. Contrary to much of what has been discussed in
the literature (e.g., refs. 6 and 34, but see ref. 35), these results
suggest that under certain circumstances, a warming climate may
be helpful to some amphibian populations, particularly those
that live in harsh conditions at the edge of their thermal toler-
ances. A reduced winter severity could benefit high and low el-
evation populations of R. luteiventris alike, although failures in
recruitment due to drying of breeding habitat may overwhelm
positive effects of warmer winters in some populations, especially
at lower elevations and latitudes.
There are several reasons that we might expect R. luteiventris
juvenile and adult survival in these systems to increase with de-
creased snowpack. At high elevations, this species spends up to 8
or 9 months overwintering in lakes and permanent ponds, where
its metabolism slows considerably. Hypoxic conditions have been
shown to be an important source of mortality for overwintering
amphibian populations in high elevation systems (18). Winters
with above-average precipitation had thicker ice covering on lakes
used by mountain yellow-legged frogs (R. muscosa) for over-
wintering, resulting in an increase in the rate of oxygen depletion
(18). Radio-tagged R. luteiventris overwintering in low elevation
ponds in eastern Oregon all remained in shallow water within 1 m
of shore, but they moved to microhabitats with higher water
temperatures or dissolved oxygen concentrations (36). Juvenile
frogs in our system may be more likely to overwinter in shallow
lakes than adult frogs, because we find them in higher concen-
trations at these water bodies during summer surveys and they
may not be able to locate larger water bodies. Overwintering
in shallow waters may expose juveniles to more hypoxic conditions
in the winter than adult frogs and likely explains the strong
negative associations among SWE and juvenile survival and
transition rates.
Years with higher snowpack tended to have snow on the ground
for a longer time periodand later into the spring. As a result, it is also
possible that high winter severity depletes the fat reserves of frogs
and leads to greater mortality rates. The later the last day of winter,
the longer frogs remain in overwintering sites, and the later insect
prey emerge. This delay in foraging and depletion of fatreserves may
contribute to the negative relationship between SWE and transition
probability to larger age classes, which we observed for juveniles.
The depletion of fat reserves may be less important for subadult and
adult frogs and may explain why we do not see a relationship be-
tween their transition rates and winter severity variables.
In all years of our study, asymptotic estimates of λwere less than
one, suggesting a declining population. This agrees with our ob-
servation that this population has generally declined over the
9 years of the study, from 997 frogs in 2000 to 700 frogs in 2008. A
decreasing population growth rate during most years is consistent
with population growth patterns measured in other amphibian
populations (37, 38). However, in contrast to our asymptotic
estimates of λ, population size fluctuated considerably among
years, with some large increases in population size between years.
Therefore, our asymptotic analyses do not capture the range of
stochastic fluctuations observed in the field. For example, it may
be that population dynamics are partly driven by intermittent
pulses of recruitment that were not captured in our 9-year study.
As such, we cannot discern whether a warming climate will actu-
ally lead to a viable population in the long term. However, this
caveat does not change the result that we observed uniformly
positive effects of climate warming.
The results of this study highlight the need for amphibian de-
cline research to focus on environmental variables affecting post-
metamorphic survival parameters and those affecting recruitment
and larval survival (23). Although repeated failures in recruitment
due to drought may critically affect the persistence of some pop-
ulations, changes in juvenile and adult survival could have a greater
effect on population growth and viability in many amphibian
species. Although we saw high annual fluctuations in the number
of egg masses laid and successful metamorphs produced in this
system (Table S1), differences in juvenile and adult survival rates,
particularly survival of juvenile frogs, drove the differences we
observed in λacross years. In our population, juveniles and adults
have greater variance in vital rates among years, as well as higher
elasticities in those rates (Table S2). Vital rates for R. luteiventris
juveniles and adults are therefore disproportionately important to
the dynamics of this population. Long-term monitoring of am-
phibian populations has typically focused on variation in re-
0
Pjj*Sj
Pjs*Sj
Pjf*Sj
Pb*Ss*Cs
0
Pss*Ss
Psf*Ss
Pb*Sf*Cs
0
0
Sf
0
S0
0
0
Fig. 4. Female-based, postbirth pulse matrix model for R. luteiventris. Each
year, average values for survival, growth, and fecundity were used to cal-
culate the asymptotic population growth rate. Average clutch size was 406
female eggs for all years based on measurements of clutch size.
y = -0.006x + 1.384
R = 0.589
0.4
0.5
0.6
0.7
0.8
0.9
1
60 70 80 90 100 110 120 130 140
Peak SWE (cm)
y = -0.010x + 2.300
R = 0.290
0.4
0.5
0.6
0.7
0.8
0.9
1
145 150 155 160 165 170 175 180 185
End of winter (days)
2001
2002
2003
2004
2005
2006
2007
2008
2008
2002
2003
2006
2004
2007
2005
2001
A
B
Fig. 5. Linear regression of annual asymptotic λvalues and winter variables,
including SWE (P= 0.021) (A) and end of winter Julian date (P= 0.185) (B).
8646
|
www.pnas.org/cgi/doi/10.1073/pnas.0912945107 McCaffery and Maxell
cruitment or breeding population size (e.g., ref. 37). Past research
has focused on the effect of weather variability on production
of juveniles, which can depend heavily on rainfall in some regions
(39). However, it is less clear to what extent this measured varia-
tion in annual recruitment affects long-term population dynamics
in these systems, because recruitment dynamics are rarely placed
in a population-level context (but see refs. 23 and 40).
Although there was variation in annual recruitment, we never
witnessed complete failures in recruitment, even in the driest years.
Every year, individuals breed in both ephemeral and permanent
water bodies. On cool and wet years, most successful recruitment
comes from ephemeral ponds, which persist throughout the grow-
ing season, whereas tadpoles from colder permanent ponds do not
have enough time to reach metamorphosis. However, in hot, dry
years the majority of recruitment comes from permanent ponds,
because ephemeral ponds dry up before animals are able to suc-
cessfully metamorphose. Therefore, it may be that frogs in our
study system have the ability to compensate recruitment in dry years
through the diversity of breeding sites available across this glacial
basin. The presence of a diversity of unimpacted breeding sites in
these types of systems may be important for amphibian populations
to persist in the face of climate change. In high-elevation systems
where amphibians are negatively impacted by introduced fish (e.g.,
ref. 31), reduction of hydroperiod in ephemeral, fishless breeding
sites may be much more important to species persistence. Because
this system is fishless, amphibians are able to breed and forage in
a greater diversity of habitats than in areas where fish and other
predators restrict populations to ephemeral ponds and wetlands.
Our analyses suggest that for amphibians and other ectotherms
living at temperatures below their physiological optima, such as
those found at high elevations or latitudes, global warming may
have neutral or positive impacts, at least under some climate
change scenarios (41). However, outcomes will surely depend on
how quickly and to what extent the climate changes. It is possible
that changes in temperature and precipitation may only have
benefits to a certain point, after which habitat will be negatively
affected, possibly by changing community structure or reducing
heterogeneity of breeding sites. At the extreme, if ongoing climate
change eventually makes high elevation and latitude sites unsuit-
able, species may have no other refugia to move to (42). Both the
positive and negative impacts of climate change on species will
need to be reassessed as climate change progresses to determine
whether such a tipping point will be reached.
We have shown how individual amphibian vital rates and as-
ymptotic population growth rates vary with summer and winter
climate variables in a high elevation population of the Columbia
spotted frog. These results unambiguously demonstrate that
earlier ending winters with lower snowpack in this system lead to
higher survival rates, higher probabilities of breeding, and higher
population viability. Most research on amphibian declines
assumes that climate change will have negative impacts on al-
ready vulnerable species, yet we show that this may not be the
case for alpine and boreal amphibian populations currently
persisting in harsh environments. This provides a unique per-
spective to the role of climate change in amphibian declines in
temperate ecosystems. Previous shifts in climate have had dra-
matic effects on the distribution, genetics, and ecology of nu-
merous temperate amphibian species (43–45). The impacts of
ongoing climate change will vary across the globe, and this is
likely to increase the viability of some species while being det-
rimental to others. To best conserve viable populations, promote
gene flow, and allow for shifts in geographic range in the face of
climate change, developing a more comprehensive un-
derstanding of how climate change is expected to benefit or harm
amphibian populations at different latitudes and elevations is
imperative. Ultimately, research that simultaneously evaluates
the effects of climate change on multiple vital rates is key to
disentangling which population processes may be affected by
a warming climate.
Materials and Methods
Population Surveys. We collected demographic data from 2000 to 2008. We
conducted systematic searches for egg masses in late spring. We searched all
shallow water environments for egg masses over several weeks and recorded
countsof egg masses at eachwater body. We calculatedmultiyearestimates and
variancesin clutch size for R. luteiven tris at mostof the breeding ponds by using
volumetric displacement (46–48). During the egg mass surveys, we also mea-
suredthe snout-vent lengthof females presentat the breeding siteto determine
the range of size of breeding females. These data were used to distinguish
breeding adult females from subadult females in our demographic analyses.
We useda robust designcapture-mark-recapture methodto monitor juvenile
and adult frogs (49). For this method, we captured animals for multiple con-
secutive secondary sessions (days) within the primary sampling period (year).
Across the secondary sessions, we assumed that the population was closed to
immigration, emigration, births, and deaths. This closure allowed population
size, corrected for capture probability, to be estimated. Between primary ses-
sions,the populationwas open to gains and losses,and survivalwas estimatedby
using the Cormack-Jolly-Seber model (50–52). We monitored three female life
stages: juveniles were frogs that were too young to be sexed, subadults were
frogs that were larger than the size at which secondary sex characteristics of
males were pre sent but smaller than the s mallest documente db reeding female,
and adults were frogs large enough to breed. During each primary sampling
period, we systematically surveyed all of the ponds and lakeshores in the basin
each yearand captured animals by handor net. We individually marked animals
by clipping unique combinations of toes by using an alphanumeric coding sys-
tem (53) and weighed and measured snout-vent length. We also recorded the
general location of all new and recaptured animals at each session.
Table 1. Model structures for assessing how climate variables relate to survival and growth
parameters
Winter variables Summer variables
Stage + (SWE) Stage + (summer length)
Stage + (end of winter) Stage + (growing degree days)
Stage + (winter length) Stage + (summer length + growing degree days)
Stage + (SWE + winter length) Stage ×(summer length)
Stage + (SWE + end of winter) Stage ×(growing degree days)
Stage ×(SWE) Stage ×(summer length + growing degree days)
Stage ×(end of winter)
Stage ×(winter length)
Stage ×(SWE + winter length)
Stage ×(SWE + end of winter)
For all models, we let pand cvary by both age class and time. We first estimated survival probabilities,
keeping growth probabilities constant, and then estimated growth probabilities by using the best model from
the survival analysis. For transition models, we examined scenarios where all transitions were different and
where probability of juvenile to male transition was equal to juvenile to subadult transition.
McCaffery and Maxell PNAS
|
May 11, 2010
|
vol. 107
|
no. 19
|
8647
ECOLOGY
Climate Data. Climate data were from the Twin Lakes SNOTEL site (www.wcc.
nrcs.usda.gov/snow), located 18 km northwest of the upper Little Rock Creek
basin at 1,950 m elevation. This SNOTEL site is located within the same
mountain range and has similar aspect and elevation to our study area. We
used data on SWE and temperature as annual covariates. SWE is the depth of
water that would theoretically result if one were to melt the entire snowpack
instantaneously and is commonly used as a measure of snowpack (e.g., ref.
21). We used the peak SWE value recorded for each year in our analyses. We
calculated the end of winter Julian date as the last day that SWE was recorded
in the spring. Then we calculated the number of consecutive days with snow
on the ground (winter length) and the number of consecutive days without
snow on the ground (summer length). Finally, we calculated the cumulative
number of days throughout the summer that the temperature was >10 °C
(www.ipm.ucdavis.edu/WEATHER/ddconcepts.html), which represented
growing degree days. The value of 10 °C was used as a relative physiological
zero for the species, temperatures below which we would not expect to see
growth. Although it is unlikely that this temperature is the true physiological
zero, which is unknown, this cutoff allows for comparison of relative growing
season length across years. We used winter length, last day of winter, and
SWE data as indices of local winter severity, and we used growin g degree days
and summer length as indices of summer intensity.
Wetested for relationshipsamong all climatevariables overthe 9 yearsof the
study by using principal components analysis. The first two axes explained 89%
of the variance. Axis 1 (56% of variance) was correlated with winter severity (r>
0.85 for all winter variables; r<0.53 for all summer variables). Axis 2 (33% of
variance) was correlated with summer variables (r>0.75 for all summer vari-
ables; r<0.40 for all winter variables). Therefore, we can clearly distinguish
winter from summer variables but cannot statistically separate individual
winter or summer variables. However, we chose to analyze all variables in-
dividually to discuss hypothetical mechanisms associated with each variable.
Breeding Probability and Prejuvenile Analysis. We estimated breeding prob-
ability as the number of egg masses deposited each year divided by the
number of adult females that year (closed population estimate from the
mark-recapture data). We estimated survival from eggs to 1 year for each year
by dividing the number of 1-year-old frogs by the number of eggs laid the
previous spring (average number of eggs per mass multiplied by the number
of egg masses deposited). We used an average of 812 eggs per mass (assumed
to be equivalent to 406 female embryos given a 1:1 sex ratio) from our clutch
size estimates for all years of the analysis. To estimate the number of 1-year-
old frogs, we isolated the capture histories for all animals 34 mm snout-vent
length and smaller each year and then calculated the abundance of those
animals by using Huggins closed population models in program MARK (32).
We used weighted linear regressions to examine the dependence of
breeding probability and first-year survival on the five climate variables
described above. Additionally, we examined relationships between number
of egg masses deposited and climate variables.
Post-Metamorph Analysis. We analyzed capture data of juveniles, subadults,
and adults in program MARK by using closed robust design, multistate mark-
recapture models (54, 55). The “states”in our models were frog life stages:
juvenile, subadult female, adult female, and adult male frogs. These models
provided parameter estimates and variances for age- and sex-specific survival,
transition rates between life stages, and population size for each year (56, 57).
Survival, transitions, and the covariates were our primary parameters of in-
terest, but we first evaluated models of capture probability (p) and recapture
probability given initial capture (c) to avoid bias and imprecision in the sur-
vival estimates (58). We evaluated the following capture probability models:
(i)pand cwere constant across years but varied among age classes, (ii )pand c
varied across years but did not vary among age classes, (iii )pand cvaried
across years and differed between juvenile and adult age classes, but not
within adult age classes, and (iv)pand cvaried across years and among all age
classes. For these models, survival and transition rates were held constant.
We modeled age-specific survival and transition probabilities to examine
their variation in relation to snowpack, winter length, end of winter, summer
length, and growing degree-days. First, we examined how each of these
variables related to survival in each age class individually. Then, we examined
whether additive models containing multiple climate variables improved the
fit. For each model structure, we estimated models where (i) the effect of
covariates was the same for each age class and (ii ) the effect of covariates
was different for each age class, resulting in a total of 16 basic models (Table
1). Once we tested these models, we also examined two additional models
where (i) male survival was equal to female survival and (ii) all adult age
classes had the same survival. For these models, we kept transition proba-
bilities constant. Once we had modeled survival, we examined how climate
variables affected growth by using the structure from the best survival
model. We examined the same suite of variables for growth as we did for
survival. In addition to the 16 models described above, we estimated models
where probability of transitioning from juvenile to subadult was equal to
transitioning to male (Table 1). We evaluated all models for relative support
and ranked them by using Akaike’s Information Criterion (AICc) (59, 60).
Demographic Analysis. We determined the demographic consequences of
changes in climate variables on the population growth rate using matrix
models (19). We constructed female-based postbirth pulse stage-structured
matrix models by using the average vital rates estimated for each year (Table
S1 and Fig. 4). Survival and transition parameters came from the best time-
dependent model ranked by using AICc. We tested four time-dependent
models where survival varied by both year and life stage: (i) survival varies in
the same way for each life stage in each year; (ii ) juvenile survival varies dif-
ferently from subadult, male, and female survival, which vary in the same way
each year; (iii) male and female survival vary in the same way, but both ju-
venile and subadult survival vary differently in each year; and (iv) all life stages
varied differently from each other in each year. The best model varied tran-
sitions for each life stage and year, but transition probability from juvenile to
subadult was equal to that for juvenile to male. Capture and recapture
probabilities varied by year and life stage for all models. Other vital rates were
calculated as described above. For each year, we calculated asymptotic λ,the
rate at which a population would increase if vital rates remained constant
over time. This metric represents an integrative measure of population per-
formance. We then used linear regression of the asymptotic lambda values
for each year against values of climate variables for each year to see whether
this integrative metric showed similar patterns to individual vital rates.
ACKNOWLEDGMENTS. We thank the many field assistants who helped
collect the demographic data used in this study and K. Griffin for assistance
with the mark-recapture analyses. We thank E. E. Crone, L. Eby, P. S. Corn,
L. S. Mills, W. H. Lowe, and two anonymous reviewers for helpful comments
and insights on analyses and drafts of the manuscript. Fieldwork was partly
supported with funding from the US Geological Survey’s Amphibian Re-
search and Monitoring Initiative through cooperative agreements among
the US Geological Survey, the University of Montana, and the Montana
Natural Heritage Program. Research was also partly supported by US Forest
Service Region 1 Inventory and Monitoring grants (to B.A.M.).
1. Alford RA, Richards SJ (1999) Global amphibian declines: A problem in applied
ecology. Annu Rev Ecol Syst 30:133–165.
2. Houlahan JE, et al. (2000) Quantitative evidence for global amphibian population
declines. Nature 404:752–755.
3. Stuart SN, et al. (2004) Status and trends of amphibian declines and extinctions
worldwide. Science 306:1783–1786.
4. Pounds JA, Crump ML (1994) Amphibian declines and climate disturbance: The case of
the golden toad and the harlequin frog. Conserv Biol 8:72–85.
5. Pounds JA, et al. (2006) Widespread amphibian extinctions from epidemic disease
driven by global warming. Nature 439:161–167.
6. Corn PS (2005) Climate change and amphibians. Anim Biodivers Conserv 28:59–67.
7. Beebee TJC (1995) Amphibian breeding and climate change. Nature 374:219–220.
8. Forchhammer MC, Post E, Stenseth NC (1998) Breeding phenology and climate.
Nature 391:29–30.
9. Blaustein AR, et al. (2001) Amphibian breeding and climate change. Conserv Biol 15:
1804–1809.
10. Corn PS (2003) Amphibian breeding and climate change: The importance of snow in
the mountains. Conserv Biol 17:622–625.
11. Reading CJ (1998) The effect of winter temperatures on the timing of breeding
activity in the common toad Bufo bufo.Oecologia 117:469–475.
12. Rohr JR, et al. (2008) Evaluating the links between climate, disease spread, and
amphibian declines. Proc Natl Acad Sci USA 105:17436–17441.
13. Lips KR, Diffendorfer J, Mendelson JR, III, Sears MW (2008) Riding the wave: Reconciling
the roles of disease and climate change in amphibian decline. PLoS Biol 6:e72.
14. Reading CJ (2007) Linking global warming to amphibian declines through its effects
on female body condition and survivorship. Oecologia 151:125–131.
15. Kiesecker JM, Blaustein AR, Belden LK (2001) Complex causes of amphibian
population declines. Nature 410:681–684.
16. Carey C (1978) Factors affecting body temperatures of toads. Oecologia 35:
197–219.
17. Wells KD (2007) The ecology and behavior of amphibians (Univ Chicago Press,
Chicago).
18. Bradford DF (1983) Winterkill, oxygen relations, and energy metabolism of
a submerged dormant amphibian, Rana muscosa.Ecology 64:1171–1183.
19. Tattersall GJ, Boutelier RG (1997) Balancing hypoxia and hypothermia in cold-
submerged frogs. J Exp Biol 200:1031–1038.
8648
|
www.pnas.org/cgi/doi/10.1073/pnas.0912945107 McCaffery and Maxell
20. Bernardo J, Spotila JR (2006) Physiological constraints on organismal response to
global warming: Mechanistic insights from clinally varying populations and
implications for assessing endangerment. Biol Lett 2:135–139.
21. Hamlet AF, Mote PW, Clark MP, Lettenmaier DP (2005) Effects of temperature and
precipitation variability on snowpack trends in the western United States. J Clim 18:
4545–4561.
22. MacCracken M, et al. (2001) Climate change impacts on the United States: the
potential consequences of climate variability and change, ed Team NAS (Cambridge
Univ Press, Cambridge, UK), pp 13–71.
23. Biek R, Funk WC, Maxell BA, Mills LS (2002) What is missing in amphibian decline
research: insights from ecological sensitivity analysis. Conserv Biol 16:728–734.
24. Caswell H (2001) Matrix population models–construction, analysis, and interpretation
(Sinauer Associates, Sunderland, MA).
25. Leftkovitch LP (1965) The study of population growth in organisms grouped by
stages. Biometrics 21:1–18.
26. Leslie PH (1945) On the use of matrices in certain population mathematics. Biometrika
33:183–212.
27. Funk WC, et al. (2005) Population structure of Columbia spotted frogs (Rana
luteiventris) is strongly affected by landscape. Mol Ecol 14:483–496.
28. Maxell BA (2009) State-wide assessment of status, predicted distribution, and
landscape-level habitat suitability of amphibians and reptiles in Montana. PhD
dissertation (University of Montana, Missoula).
29. Rollins-Smith LA, et al. (2002) Activity of antimicrobial skin peptides from ranid frogs
against Batrachochytrium dendrobatidis, the chytrid fungus associated with global
amphibian declines. Dev Comp Immunol 26:471–479.
30. Muths E, Corn PS, Pessier AP, Green DE (2003) Evidence for disease-related amphibian
decline in Colorado. Biol Conserv 110:357–365.
31. Knapp RA, Matthews KR (2000) Non-native fish introductions and the decline of the
mountain yellow-legged frog from within protected areas. Conserv Biol 14:428–438.
32. White GC, Burnham KP (1999) Program MARK: Survival estimation from populations
of marked animals. Bird Study 46 (Suppl):120–138.
33. Adam JC, Hamlet AF, Lettenmaier DP (2009) Implications of global climate change for
snowmelt hydrology in the twenty-first century. Hydrol Process 23:962–972.
34. McMenamin SK, Hadly EA, Wright CK (2008) Climatic change and wetland desiccation
cause amphibian decline in Yellowstone National Park. Proc Natl Acad Sci USA 105:
16988–16993.
35. Scherer RD , Muths E, Lambert BA (2008) Effects of weather on survival in populations
of boreal toads in Colorado. J Herpetol 42:508–517.
36. Bull EL, Hayes MP (2002) Overwintering of Columbia spotted frogs in northeastern
Oregon. Northwest Sci 76:141–147.
37. Semlitsch RD, Scott DE, Pechmann JHK, Gibbons JW (1996) Long-Term Studies of
Vertebrate Communities, edsCody ML Smallwood JA (Academic, San Diego),pp 217–248.
38. Pechmann JHK, et al. (1991) Declining amphibian populations: The problem of
separating human impacts from natural fluctuations. Science 253:892–895.
39. Daszak P, et al. (2005) Amphibian population declines at Savannah River Site are
linked to climate, not chytridiomycosis. Ecology 86:3232–3237.
40. Vonesh JR, De La Cruz O (2002) Complex life cycles and density dependence: assessing
the contribution of egg mortality to amphibian declines. Oecologia 133:325–333.
41. Deutsch CA, et al. (2008) Impacts of climate warming on terrestrial ectotherms across
latitude. Proc Natl Acad Sci USA 105:6668–6672.
42. Wake DB, Vredenburg VT (2008) Are we in the midst of the sixth mass extinction? A
view from the world of amphibians. Proc Natl Acad Sci USA 105:11466–11473.
43. Funk WC, et al. (2008) Range-wide phylogeographic analysis of the spotted frog
complex (Rana luteiventris and Rana pretiosa) in northwestern North America. Mol
Phylogenet Evol 49:198–210.
44. Goebel AM, Ranker TA, Corn PS, Olmstead RG (2009) Mitochondrial DNA evolution in
the Anaxyrus boreas species group. Mol Phylogenet Evol 50:209–225.
45. Vieites DR, Min M-S, Wake DB (2007) Rapid diversification and dispersal during
periods of global warming by plethodontid salamanders. Proc Natl Acad Sci USA 104:
19903–19907.
46. Morris RL, Tanner WW (1969) The ecology of the western spotted frog, Rana pretiosa
pretiosa Baird and Girard: A life history study. Great Basin Nat 29:45–81.
47. Corn PS, Livo LJ (1989) Leopard frog and wood frog reproduction in Colorado and
Wyoming. Northwest Nat 70:1–9.
48. Werner JK, Weaselhead J, Plummer T (1999) The accuracy of estimating eggs in
anuran egg masses using weight or volume measurements. Herpetol Rev 30:30–31.
49. Pollock KH (1982) A capture-recapture design robust to unequal probability of
capture. J Wildl Manage 46:752–757.
50. Cormack RM (1964) Estimates of survival from the sighting of marked animals.
Biometrika 51:429–438.
51. Jolly GM (1965) Explicit estimates from capture-recapture data with both death and
immigration stochastic model. Biometrika 52:225–247.
52. Seber GAF (1965) A note on the multiple recapture census. Biometrika 52:249–259.
53. Waichman AV (1992) An alphanumeric code for toe clipping amphibians and reptiles.
Herpetol Rev 23:19–21.
54. Hestbeck JB, Nichols JD, Malecki RA (1991) Estimates of movement and site fidelity
using mark-resight data of wintering Canada geese. Ecology 72:523–533.
55. White GC, Kendall BE, Barker RJ (2006) Multistate survival models and their
extensions in Program MARK. J Wildl Manage 70:1521–1529.
56. Morris WF, Doak DF (2002) Quantitative Conservation Biology (Sinauer Associates,
Sunderland, MA).
57. Kendall BE (1998) Estimating the magnitude of environmental stochasticity in
survivorship data. Ecol Appl 8:184–193.
58. Lebreton JD, Burnham KP, Clobert J, Anderson DR (1992) Modeling survival and
testing biological hypotheses using marked animals: a unified approach with case
studies. Ecol Monogr 62:67–118.
59. Akaike H (1973) Second International Symposium on Information Theory, eds Csaki F
Petrov BN (Akademiai Kiado, Budapest).
60. Burnham KP, Anderson DR (2002) Model Selection and Inference: A Practical
Information-Theoretic Approach (Springer, New York).
McCaffery and Maxell PNAS
|
May 11, 2010
|
vol. 107
|
no. 19
|
8649
ECOLOGY
