Content uploaded by Brian Todd
Author content
All content in this area was uploaded by Brian Todd on Oct 17, 2017
Content may be subject to copyright.
Remarkable Amphibian Biomass and Abundance in an
Isolated Wetland: Implications for Wetland
Conservation
J. WHITFIELD GIBBONS,∗†∗∗∗ CHRISTOPHER T. WINNE,∗† DAVID E. SCOTT,∗JOHN D.
WILLSON,∗† XAVIER GLAUDAS‡, KIMBERLY M. ANDREWS,∗† BRIAN D. TODD,∗† LUKE A.
FEDEWA§, LUCAS WILKINSON,∗RIA N. TSALIAGOS,∗∗ STEVEN J. HARPER,∗† JUDITH L. GREENE,∗
TRACEY D. TUBERVILLE,∗† BRIAN S. METTS,∗† MICHAEL E. DORCAS††, JOHN P. NESTOR,∗
CAMERON A. YOUNG,∗† TOM AKRE,∗ROBERT N. REED‡‡, KURT A. BUHLMANN,∗
JASON NORMAN,∗DEAN A. CROSHAW,∗§§ CRIS HAGEN,∗AND BETSIE B. ROTHERMEL∗
∗Savannah River Ecology Laboratory, University of Georgia, Drawer E, Aiken, SC 29802, U.S.A.
†Institute of Ecology, University of Georgia, Athens, GA 30608, U.S.A.
‡University of Nevada, Las Vegas, 4505 Maryland Parkway, Las Vegas, NV 89154–4004, U.S.A.
§Partners in Amphibian and Reptile Conservation, 2221 West Greenway Road, Phoenix, AZ 85023, U.S.A.
∗∗Wayland Baptist University, 5530 E. Northern Lights Boulevard, Suite 24, Anchorage, AK 99504, U.S.A.
††Department of Biology, Davidson College, Davidson, NC 28035–7118, U.S.A.
‡‡Department of Biology, Southern Utah University, Cedar City, UT 84720, U.S.A.
§§Department of Biological Sciences, University of New Orleans, New Orleans, LA 70148–0001, U.S.A.
Abstract: Despite the continuing loss of wetland habitats and associated declines in amphibian populations,
attempts to translate wetland losses into measurable losses to ecosystems have been lacking. We estimated
the potential productivity from the amphibian community that would be compromised by the loss of a single
isolated wetland that has been protected from most industrial, agricultural, and urban impacts for the past
54 years. We used a continuous drift fence at Ellenton Bay, a 10-ha freshwater wetland on the Savannah River
Site, near Aiken, South Carolina (U.S.A.), to sample all amphibians for 1 year following a prolonged drought.
Despite intensive agricultural use of the land surrounding Ellenton Bay prior to 1951, we documented 24
species and remarkably high numbers and biomass of juvenile amphibians (>360,000 individuals; >1,400
kg) produced during one breeding season. Anurans (17 species) were more abundant than salamanders
(7 species), comprising 96.4% of individual captures. Most (95.9%) of the amphibian biomass came from
232095 individuals of a single species of anuran (southern leopard frog [Rana sphenocephala]). Our results
revealed the resilience of an amphibian community to natural stressors and historical habitat alteration and
the potential magnitude of biomass and energy transfer from isolated wetlands to surrounding terrestrial
habitat. We attributed the postdrought success of amphibians to a combination of adult longevity (often >5
years), a reduction in predator abundance, and an abundance of larval food resources. Likewise, the increase
of forest cover around Ellenton Bay from <20% in 1951 to >60% in 2001 probably contributed to the long-
term persistence of amphibians at this site. Our findings provide an optimistic counterpoint to the issue of the
global decline of biological diversity by demonstrating that conservation efforts can mitigate historical habitat
degradation.
Keywords: amphibian decline, biodiversity, drought, land use, wetland recovery
∗∗∗email gibbons@srel.edu
Paper submitted June 7, 2005; revised manuscript accepted November 7, 2005.
1457
Conservation Biology Volume 20, No. 5, 1457–1465
C
2006 Society for Conservation Biology
DOI: 10.1111/j.1523-1739.2006.00443.x
1458 Amphibian Biomass and Abundance Gibbons et al.
Biomasa y Abundancia de Anfibios Extraordinaria en un Humedal Aislado: Implicaciones para la Conservaci´on de
Humedales
Resumen: A pesar de la p´
erdida de h´
abitats de humedales y las declinaciones asociadas de poblaciones de
anfibios, se han realizado pocos intentos para traducir las p´
erdidas de humedales en p´
erdidas mensurables
en los ecosistemas. Estimamos la productividad potencial de la comunidad de anfibios que se afectar´
ıa por
la p´
erdida de un humedal aislado que ha estado protegido de los impactos industriales, agr´
ıcolas y urbanos
durante los ´
ultimos 54 a˜
nos. Utilizamos un cerco de desv´
ıo en la Bah´
ıa Ellentonn, un humedal dulceacu´
ıcola de
10 ha en el R´
ıo Savannah, cerca de Aiken, Carolina del Sur (E.U.A.), para muestrear todos los anfibios durante
1a
˜
no despu´
es de una sequ´
ıa prolongada. A pesar del intensivo uso agr´
ıcola del suelo alrededor de la Bah´
ıa
Ellenton antes de 1951, documentamos 24 especies y n´
umeros y biomasa de anfibios juveniles notablemente
altos (>360,000 individuos; >1,400 kg) en una temporada reproductiva. Los anuros (17 especies) fueron m´
as
abundantes que las salamandras (7 especies), y comprendieron 96.4% de las capturas individuales. La mayor
parte (95.9%) de la biomasa provino de 232095 individuos de una sola especie de anuro (Rana sphenocephala).
Nuestros resultados revelaron que la resiliencia de la comunidad de anfibios a los estresantes naturales y a
la alteraci´
on hist´
orica del h´
abitat y la magnitud potencial de la transferencia de biomasa y energ´
ıa desde los
humedales aislados hacia el h´
abitat terrestre circundante. Atribuimos el ´
exito post-sequ´
ıa de los anfibios a una
combinaci´
on de longevidad de adultos (a menudo >5a
˜
nos), la reducci´
on de la abundancia de depredadores
y la abundancia de recursos alimenticios para las larvas. Asimismo, el incremento de la cobertura forestal
alrededor de la Bah´
ıa Ellerton de <20% en 1951 a >60% en 2001 probablemente contribuy´
o a la persistencia
de los anfibios a largo plazo en este sitio. Nuestros hallazgos proporcionan un contrapunto optimista al tema
de la declinaci´
on global de la diversidad biol´
ogica al demostrar que los esfuerzos de conservaci´
on pueden
mitigar a la degradaci´
on hist´
orica del h´
abitat.
Palabras Clave: biodiversidad,declinaci´on de anfibios, recuperaci´on de humedales sequ´ıa, uso de suelo
Introduction
Isolated wetlands, which support high species diversities
and serve as essential habitat for many groups of organ-
isms, are under increased threat of destruction. In the
United States, the ongoing decline in numbers of iso-
lated and other palustrine wetlands due to agricultural
and commercial development since European settlement
is well documented (Dahl 1990, 2000). At least 87% of
the original 1.2 million ha of shrub bog pocosins (see de-
scription in Sharitz & Gibbons 1982) in the southeastern
United States have been destroyed or altered (Richardson
1983). Similarly, 69% of playas >4 ha in the Southern Great
Plains have been modified by cultivation, grazing, and
other activities (Guthery & Bryant 1982). The Solid Waste
Agency of Northern Cook County v. U.S. Army Corps of
Engineers (2000) decision by the U.S. Supreme Court has
weakened federal jurisdiction over isolated wetlands, ef-
fectively leaving freshwater wetlands unprotected when
they lack a permanent surface-water connection to navi-
gable waterways and have no link to interstate commerce
(Zedler et al. 2001; Downing et al. 2003). The loss and
degradation of numerous wetlands have presumably re-
sulted in a concomitant loss of species abundance and
diversity, with ramifications for ecosystem functioning.
Despite the ecological value of isolated wetlands to re-
gional biodiversity (Whigham 1999) and their importance
for maintaining metapopulation connectivity for semi-
aquatic species (Gibbs 1993; Semlitsch & Bodie 1998),
few researchers have quantified the productivity achiev-
able within a taxonomic group from a single isolated
wetland over a well-defined time frame. Moreover, al-
though surrounding terrestrial habitats are critical for
the integrity of wetland ecosystems (Gibbons 2003), few
researchers have enumerated the potential transfer of
biomass and energy from wetlands to surrounding ter-
restrial ecosystems.
Our study was undertaken to document the potential
contribution of amphibians to secondary production and
to the transfer of matter and energy between aquatic and
terrestrial habitats. Although the global decline of am-
phibians has been confirmed repeatedly (Stokstad 2004;
Stuart et al. 2004), attempts to translate wetland losses
into tangible effects on ecosystems due to the loss of
amphibian productivity have been lacking. Our results
demonstrate the levels of amphibian abundance and di-
versity that contribute to habitat interconnectivity and
that could potentially be lost by the elimination of a single
isolated wetland. We also addressed the critical question
of how rapidly and successfully amphibian populations
inhabiting ephemeral wetlands in a historically agricul-
tural landscape can recover from multiyear droughts.
Methods
Study Site
Ellenton Bay (Gibbons 1990) is an isolated freshwater wet-
land, typical of Carolina bays (see descriptions in Sharitz
& Gibbons 1982; Sharitz 2003), and within a 202-ha
Conservation Biology
Volume 20, No. 5, October 2006
Gibbons et al. Amphibian Biomass and Abundance 1459
Figure1. Aerial photos from (a) 1951 and (b) 2001 of land cover surrounding Ellenton Bay (arrows) and (c)
relationship between distance from Ellenton Bay and the proportion of the landscape within 2 km of the periphery
of Ellenton Bay classified as upland forest in 1951 and in 2001. In 1951 Ellenton Bay was within a 202-ha
agricultural field planted in cotton and corn. By 2001 natural succession had restored the landscape to primarily
pine and mixed pine–hardwood forest.
formerly agricultural field that was planted in cotton,
peanuts, and corn from the 1870s through 1951 (Fig. 1;
Davis & Janecek 1997). The fish-free wetland is in South
Carolina on the U.S. Department of Energy’s (DOE) Savan-
nah River Site (SRS), a national environmental research
park (Shearer & Frazer 1997). Wetlands on the SRS have
been protected since 1951 from most of the environmen-
tal perturbations typically resulting from agricultural, ur-
ban, and industrial alterations in the region. Ellenton Bay
and surrounding uplands have received special protec-
tion as part of the DOE Set-Aside Program, which was
established to protect habitats and facilitate long-term re-
search. No herpetofaunal diversity or population data are
available for Ellenton Bay and surrounding areas before
establishment of the SRS. Because of the historical empha-
sis on clearing and preparing lands for agricultural pur-
poses, we presume that species diversity and abundance
at the wetland in the first half of the twentieth century
were reduced from historical levels. Nonetheless, long-
term (since 1968) herpetological studies conducted at
Ellenton Bay suggest that the amphibian species richness
characteristic of the region (Gibbons & Semlitsch 1991)
had recovered within 25 years. After 1951, Ellenton Bay
and other wetlands on the SRS returned to more natu-
ral conditions as a consequence of secondary succession
coupled with limited forest management activities.
The closest isolated wetland to Ellenton Bay is ∼200
m away, and Ellenton Bay has the longest hydroperiod
of the nonpermanent wetlands in the region. The only
wetland within 1.4 km of Ellenton Bay that holds water
during multiyear droughts is a small, human-made pond
∼0.5 km from the bay. The areal extent of the Ellenton Bay
basin when full is approximately 10 ha. A dike 5 to 6 m
wide divides the bay completely and creates two distinct
aquatic areas when water depth is <1.5 m. Water surface
area and depth are extremely variable, with a maximum
depth of approximately 2 m. During periods of drought,
subsurface moisture normally remains beneath the thick
organic crust that covers the entire basin up to 0.5 m
deep (Gibbons 1990). Parts of the basin remain mucky
during some dry periods, with up to 1 ha of viscous mud
surrounding small areas of open water, but no standing
water remained during the 2000–2003 drought. Two ma-
jor multiyear droughts have occurred at Ellenton Bay dur-
ing the last 29 years (Willson et al., 2006), the first from
October 1987 through August 1990, and the second from
August 2000 to February 2003. We initiated the current
study in February 2003 at the end of the second drought.
Historical Changes in Land Use
We used object-based classification of aerial photographs
to quantify the change in land cover between 1951 and
2001. We classified groups of pixels, rather than individ-
ual pixels, which allowed us to incorporate both spec-
tral (i.e., red–green–blue [RGB] values) and contextual
information (i.e., spatial configuration) of objects (Walter
2004). To define our area of interest, we used a bounding
rectangle that fully encompassed the 2-km radius from the
edge of Ellenton Bay. Panchromatic aerial photographs,
with an on-ground resolution of 2 m, were taken in 1951
in conjunction with the establishment of the DOE facil-
ity. False-color, infrared, aerial photographs were taken in
2001. We resampled from 0.3 to2mtoprovide a consis-
tent on-ground resolution between time periods. We used
filters and advanced learning algorithms incorporated in
Feature Analyst (version 3.4, Visual Learning Systems, Mis-
soula, Montana), an extension of ArcGIS (version 8.3,
Conservation Biology
Volume 20, No. 5, October 2006
1460 Amphibian Biomass and Abundance Gibbons et al.
Environmental Systems Research Institute [ESRI], Red-
lands, California), to conduct supervised classifications.
We segmented images and extracted general types of land
cover that were easily recognized (i.e., Anderson level 1
classes of urban, agricultural, forest, water, wetland, and
barren) for both time periods. We generated summary
statistics of classified maps for discrete 200-m-wide buffer
intervals radiating out from the perimeter of Ellenton Bay.
Collection Techniques
In February 2003, we re-erected a terrestrial drift fence
(Gibbons & Semlitsch 1981; Semlitsch et al. 1996) that
had been used in studies at Ellenton Bay for all or part
of 19 of the 28 years from 1968 to 1994 (Gibbons 1990;
Seigel et al. 1995). The continuous drift fence completely
encircled the wetland and was equipped with pitfall and
funnel traps, which allowed us to capture most of the
amphibians entering and exiting the wetland. Serendipi-
tously, we began monitoring the fence and traps 3 days
before the start of a long period of rain that began to
refill the wetland, after 2.5 years of drought. We moni-
tored the fence from 1 February 2003 to 31 January 2004
and therefore were able to document the abundance and
productivity of amphibians throughout the first wet year
following the drought.
The drift fence was constructed of aluminum flashing
(1230 m long, 40 cm high) and buried several cm into
the soil (Gibbons & Semlitsch 1981). The distance of the
fence from the margin of the water varied with water level
but was <10 m in many places when the bay reached its
maximum water level in August 2003. We placed 164
traps in pairs on opposite sides of the fence, with half of
the traps on each side of the fence, allowing individual
amphibians to be categorized as entering or leaving the
bay (Gibbons & Semlitsch 1981). Of these, 41 pairs of
19-L pitfall traps (plastic buckets), spaced approximately
every 30 m along the fence, were in place on 1 February
2003. On 24 February 2003, we installed 21 pairs of 2.3-L
pitfall traps (metal coffee cans) between every other pair
of buckets along the fence (60 m apart). On 27 February
2003, we installed 20 pairs of wooden box funnel traps
(Himes 2000; Zappalorti & Torocco 2002) along the fence
(60 m apart) between bucket pairs where cans had not
been placed. Thus, bucket pairs were followed alternately
by cans and funnel traps.
Pitfall and funnel traps were checked a minimum of
once daily (0700–0900 hours). During warm months,
traps were checked again in the late afternoon (1700–
2000 hours). Sponges were placed in the bottom of buck-
ets and cans, providing moisture during dry conditions
or a “raft” if water collected in buckets between checks;
however, standing water was bailed from buckets and
cans daily as needed. We classified captured amphib-
ians as recently metamorphosed individuals or adults (see
“Productivity and Biomass”) and released them approxi-
mately 10 m away on the opposite side of the fence.
During peak emigrations of recently metamorphosed
amphibians (e.g., ∼7,000 to 41,000 individuals per
night), special procedures were needed to reduce mortal-
ity associated with predation and overcrowding in traps.
Box traps were not deployed at night, and in addition
to morning and afternoon checks, all other traps were
checked multiple times during the night to aid emigra-
tion of recently metamorphosed individuals. On a few
occasions, we did not use box traps during the day to re-
duce the possibility of heat-related mortality of amphib-
ians. During peak emigrations involving thousands of ani-
mals, counting all individuals of some species would have
resulted in unnecessary mortality due to prolonged reten-
tion of animals in traps. Consequently, we estimated the
number of captured animals by counting the number of
animals contained in a handful or one sweep of a dipnet
and the number necessary to empty the trap. During mass
emigrations, animals were released approximately 30 m
on the outside of the fence to discourage immediate re-
capture. Recently metamorphosed amphibians captured
entering the wetland during peak emigrations were as-
sumed to have been emigrants that had been captured
and released earlier, rather than being immigrants from
other wetlands; therefore, they were re-released on the
outside of the fence and not included in capture totals.
Productivity and Biomass
To determine amphibian productivity at Ellenton Bay, we
estimated the number of emigrating young of year of all
species. Because the wetland was dry during the previous
2.5 years, recently metamorphosed amphibians captured
leaving Ellenton Bay during this study could be catego-
rized unambiguously as having been produced during the
2003 breeding season. One or more of three criteria were
used to classify amphibians emigrating from the wetland
as recently metamorphosed: (1) similar in size to pub-
lished data on size at metamorphosis for the particular
species from this region, (2) incomplete resorption of the
tail in anurans or of the gills in salamanders, (3) species-
specific attributes indicating recent metamorphosis (e.g.,
a distinct ventral stripe in A. talpoideum; scars at base of
forearms in anurans).
To estimate body size at metamorphosis, we haphaz-
ardly collected subsamples of newly metamorphosed
individuals (n=1–148) for focal species and measured
snout-vent length (SVL; ±0.5 mm) and wet body mass
(±0.01 g). We used the mean wet mass to estimate the
total biomass for each species produced at Ellenton Bay
during 2003. Individuals that were part of the 2003 juve-
nile cohort but were older juveniles at the time of capture
(i.e., likely had experienced significant postmetamorphic
growth while still inside the fence) were not used for
Conservation Biology
Volume 20, No. 5, October 2006
Gibbons et al. Amphibian Biomass and Abundance 1461
estimates of body size at metamorphosis, but their counts
were used in estimating total biomass.
At Ellenton Bay, the mole salamander (A. talpoideum)
is the only species that undergoes facultative paedomor-
phosis (Patterson 1978; Semlitsch et al. 1990). Thus, we
subsampled metamorphosed individuals of this species in
two different seasons (spring and fall) to generate sepa-
rate estimates of body size at metamorphosis and biomass.
By January 2004 it became difficult to distinguish whether
emigrating animals were large, newly metamorphosed in-
dividuals from breeding that occurred in 2003 (and there-
fore should have been tallied for the biomass estimate) or
were postbreeding adults from the 2003–2004 season. Al-
though some of these salamanders were no doubt part of
the 2003 cohort, we did not count them as such. Conse-
quently, our total biomass estimate for A. talpoideum is
a highly conservative one.
To convert amphibian abundance to density estimates,
we first had to calculate the maximum area of the wetland.
We used global positioning system technology (Trimble
Pro-XR, Sunnyvale, California, with submeter accuracy) to
map the perimeter of the Ellenton Bay basin. We defined
the perimeter of the maximum waterline as the bound-
ary between emergent grasses (predominantly Panicum
spp.) in the bay’s basin and the surrounding pine (Pinus
spp.) forest. We used ArcView (version 3.3, ESRI) to cal-
culate the area of the resulting polygon.
Table 1. Total number of young-of-year amphibian emigrants at a single isolated wetland, Ellenton Bay, in South Carolina (U.S.A.).a
Total Aquatic
biomass Number of Mean density Production
Scientific name Common name (kg) individuals mass (g) (animals/ha) (kg/ha/year)
Salamanders
Ambystoma opacum marbled salamander 0.35 104 3.33 11 0.04
Ambystoma tigrinum tiger salamander 25 1,171 21.21 125 3
Ambystoma talpoideum mole salamander—late 14 2,412 5.91 257 2
summer metamorphosis
Ambystoma talpoideum mole salamander—spring 29 6,046 4.87 643 3
metamorphosis
total 68.35 9,733 — 1036 8.04
Anurans
Hyla chrysoscelisbgray treefrog — 1 — 0.1—
Hyla squirellabsquirrel treefrog — 10 — 1 —
Hyla cinereabgreen treefrog — 15 — 2 —
Acris gryllus southern cricket frog — 56 — 6 —
Scaphiopus holbrookii eastern spadefoot toad 0.17 316 0.55 34 —
Pseudacris crucifer spring peeper 0.89 1,970 0.45 210 0.09
Rana clamitans green frog (bronze frog) 1 216 5.63 23 0.13
Hyla gratiosabbarking treefrog 1 362 3.88 39 0.15
Pseudacris ornata ornate chorus frog 4 3,126 1.20 333 0.40
Bufo terrestris southern toad 48 115,056 0.42 12,240 5
Rana sphenocephalabsouthern leopard frog 1307 232,095 5.63 24,691 139
total 1421 353,223 — 37,577 151
Amphibian total 1490 362,956 — 38,612 159
aSecondary production during the 1-year study period was estimated as the number of young of year captured leaving the wetland, minus the
number of young of year captured entering the wetland.
bTotal captures were likely significantly underestimated due to the species’ ability to climb or jump over the drift fence. Estimates of biomass
are the number of young-of-year emigrants multiplied by the mean individual mass of 1–135 haphazardly selected individuals. The individual
mass used for R. clamitans was estimated from the mass of similarly sized R. sphenocephala.
Results
Historical Changes in Land Use
Historical accounts from inhabitants of the area suggest
that intensive agriculture began in the 1800s, and agri-
cultural alteration of the landscape was conspicuous in
1951, at which time forested habitat comprised <20% of
the area within 1 km of Ellenton Bay (Fig. 1). The last row
of crops—primarily corn, cotton, and peanuts—were har-
vested from fields adjacent to Ellenton Bay in 1951 (Davis
& Janecek 1997). In 1957 pine trees were planted within
80 m of the south end of the wetland, and natural estab-
lishment of pines adjacent to the bay began to occur by
the mid-1960s. Since that time, Ellenton Bay and most of
the surrounding fields have undergone natural vegetation
succession, with forest coverage within 1 km of the bay
increasing to 60–75% by 2001 (Fig. 1).
Productivity and Biomass
During the single year of drift-fence sampling at Ellen-
ton Bay, we captured 408,220 amphibians representing
24 species. Of the individual captures, 96.4% were anu-
rans (17 species) and the remainder were salamanders (7
species). Most (n=21) of the amphibian species at El-
lenton Bay were captured by April 2003, and all had been
captured by 19 August 2003. At least 362,956 recently
Conservation Biology
Volume 20, No. 5, October 2006
1462 Amphibian Biomass and Abundance Gibbons et al.
metamorphosed amphibians produced at Ellenton Bay
emigrated to the surrounding terrestrial habitat (Table
1). Most (95.9%) of the amphibian biomass came from
232095 individuals (1307 kg) of a single species of anuran
(southern leopard frog [R. sphenocephala]; Table 1). Two
species of salamanders (mole salamander [Ambystoma
talpoideum] and tiger salamander [A. tigrinum]) and one
species of anuran (southern toad [Bufo terrestris]) each
produced more than 24 kg of biomass during the year and
collectively produced more than 115 kg (Table 1). Mini-
mum estimates of larval amphibian density and biomass
for the aquatic portion of Ellenton Bay were 38612 indi-
viduals/ha and 159 kg/ha/year, based on the maximum
area inundated and numbers of emigrating young of year
(Table 1). Productivity of R. sphenocephala alone was at
least 139 kg/ha/year (Table 1). Young-of-year densities (in-
dividuals per ha) for the two most abundant species were
24,691 (R. sphenocephala) and 12240 (B. terrestris;Table
1). The productivity levels we observed for most amphib-
ian species are likely substantial underestimates, because
individuals were recorded only terrestrially when they
departed from the aquatic area of the wetland.
Discussion
Our findings on biomass and density demonstrate that
amphibians are key components of wetland ecosystems
and can potentially supply an appreciable portion of the
energy transferred between aquatic and terrestrial habi-
tats. Although high, our estimates of individual numbers
and biomass of amphibians from a single natural wetland
are conservative because of sampling protocols. For ex-
ample, many individuals of some amphibian species did
not leave the aquatic habitat during the study owing to de-
layed metamorphosis. Some species also tend to remain
within or near aquatic areas following metamorphosis,
making them less likely to emigrate as far as the drift
fence. In addition, some climbing species of anurans (e.g.,
hylid treefrogs [Dodd 1991]; Table 1) are capable of sur-
mounting drift fences, leading to an underestimation of
abundance and young-of-year biomass for these species.
Thus, the actual contribution of the amphibian commu-
nity to secondary productivity at Ellenton Bay during a
single year was appreciably higher than observed.
Previous estimates of wetland productivity have typi-
cally focused on primary production. Estimates of wet-
land primary productivity for vascular plants and algae
(e.g., Barker & Fulton 1979; Neckles 1984; Hooper &
Robinson 1976; Brinson et al. 1981) vary by an order
of magnitude, depending on factors such as hydrology,
timing and duration of inundation, latitude, and domi-
nant plant species (Brinson et al. 1981). Net primary pro-
duction of three depressional wetlands in South Carolina
ranged from 564 to 774 g/m2/year as measured by stem
production and litterfall (Busbee et al. 2003).
Secondary production in freshwater wetlands has been
measured less often than primary production and most es-
timates are of invertebrate production (e.g., White 1985;
Plante & Downing 1989; Leeper & Taylor 1998; Taylor
et al. 1989). Information on productivity or biomass of
vertebrates in aquatic systems, other than amphibians, is
limited and includes studies on fish (e.g., Lawler et al.
1974), turtles (e.g., Iverson 1982; Congdon et al. 1986),
and snakes (e.g., Godley 1980; Shine 1986). We did not
determine terrestrial densities of the Ellenton Bay amphib-
ian species. However, our density estimate of 38612 indi-
viduals/ha in the aquatic habitat is consistent with results
of previous studies indicating that amphibians are among
the most abundant vertebrates in aquatic and terrestrial
systems. For example, terrestrial plethodontid salaman-
ders reached densities of 2000–2500 individuals/ha in a
New Hampshire forest (Burton & Likens 1975) and at
least 18486 individuals/ha in a streamside habitat in the
Southern Appalachians (Petranka & Murray 2001). Maxi-
mum biomass and numbers of larval tiger salamanders (A.
tigrinum) were estimated to be 180 kg/ha and 5000 indi-
viduals/ha in prairie ponds of the western United States
(Deutschman & Peterka 1988). Our productivity estimate
(159 kg/ha/year) of amphibians that successfully emi-
grated from the aquatic habitat demonstrates that isolated
wetlands contributed substantially to the overall produc-
tivity of the surrounding landscape.
The numbers of individuals and species of amphibians
we observed, considering both the immediate drought
and history of intensive agricultural use of the Ellenton
Bay study site, demonstrate that wetland functions can re-
cover from both natural and anthropogenic disturbances
under some circumstances. During drought years in the
Coastal Plain of the southeastern United States, when
isolated wetlands either do not fill or have extremely
short hydroperiods, amphibian populations can expe-
rience complete reproductive failure (i.e., zero recruit-
ment). Ellenton Bay presumably had minimal amphibian
productivity during the 3 years preceding the 2003 study,
based on known responses of amphibians to drought and
extrapolation from observations at Rainbow Bay, a Car-
olina bay wetland located 11.5 km from Ellenton Bay
(Semlitsch et al. 1996). Continuous long-term data (1978–
2004) for 13 amphibian species at Rainbow Bay indicate
that all species suffered virtually complete reproductive
failures for 4 of the first 16 years of study owing to short-
ened hydroperiods (Semlitsch et al. 1996). From 2000
to 2003, when both the Rainbow Bay and Ellenton Bay
breeding sites remained dry for all or most of each year,
limited or no juvenile recruitment occurred at Rainbow
Bay (D.E.S., J.L.G. & B.S.M., unpublished data).
The continued presence and productivity of amphib-
ians at Ellenton Bay in 2003–2004, despite prolonged
drought, was due in part to an extensive storage compo-
nent of long-lived individuals in habitats outside the wet-
land and a reduction in predators and increase in prey
Conservation Biology
Volume 20, No. 5, October 2006
Gibbons et al. Amphibian Biomass and Abundance 1463
within the wetland. Evidence suggests that most of the
breeding adults that repopulated Ellenton Bay in 2003
were individuals that had persisted in the surrounding
terrestrial habitat rather than being dispersers from other
wetlands. A capture–recapture study at Rainbow Bay doc-
umented longevities of >5–10 years for individuals of
at least seven of the species that occur at Ellenton Bay,
and some of these species (e.g., Ambystoma spp., B. ter-
restris) are known to abide drought in terrestrial habitats
adjacent to the wetland (D.E.S., unpublished data).
Although the persistence of many amphibian popula-
tions at the landscape level depends on dispersal and re-
colonization from other wetlands (Semlitsch 2000), most
amphibians arriving at Ellenton Bay did not immigrate
from the direction of the only nearby permanent water
body, but instead arrived at the drift fence from all direc-
tions. All ephemeral wetlands within 2 km of Ellenton Bay
have shorter hydroperiods and would not have served as a
refuge for amphibians during the preceding drought. Fol-
lowing droughts, when adult amphibians return to Ellen-
ton Bay to breed, their larvae benefit both from a reduced
abundance of predators (e.g., aquatic snakes, turtles, and
predatory insects; Gibbons et al. 1983; Seigel et al. 1995;
Taylor et al. 1999; Willson et al. 2006) and from enhanced
nutrient levels that support primary production (i.e., al-
gae) and invertebrate prey (e.g., copepods and cladocer-
ans; Taylor et al. 1988; Taylor & Mahoney 1990). Such
years of strong recruitment are in effect “stored” in the
population because the persistence and staggered repro-
duction of long-lived adults buffer against fluctuations in
juvenile recruitment from the aquatic stage (Warner &
Chesson 1985; Taylor et al. 2006).
Studies of the relationship between landscape charac-
teristics and amphibian distribution patterns have gen-
erally found positive associations between amphibians
and forested habitat (e.g., Hecnar & M’Closkey 1996;
Gibbs 1998; Willson & Dorcas 2003; Porej et al. 2004).
Assessing the effects of previous agricultural activities
on amphibians is often equivocal, however, because of
limited historical information and because associations
between amphibians and agriculture vary regionally ac-
cording to landscape context and species composition.
Some researchers have detected decreases in amphibian
diversity in landscapes with intensive agriculture (Bonin
et al. 1997; Hecnar 1997). However, results of a study
in Iowa and Wisconsin showed little effect or slightly
positive effects of agriculture on amphibian richness and
abundance in Wisconsin and more negative associations
in Iowa (Knutson et al. 1999). Both the extent and prox-
imity of forested lands to wetland breeding habitat in-
fluence amphibian species occurrence and abundance.
Results of a study of forest extent and adjacency around
116 breeding ponds in Maine showed that five of nine
amphibian species (three of which were woodland sala-
manders) were positively associated with forest area in
the surrounding landscape (Guerry & Hunter 2002). Sim-
ilarly, Gibbs (1998) noted that two species of woodland
salamanders do not persist once forest cover is reduced
below a threshold level (30–50%).
In 1951 forest comprised <20% of the area within 1 km
of Ellenton Bay, and it is likely that the salamander com-
ponent of the amphibian fauna (Ambystoma opacum
[marbled salamander], A. talpoideum,A. tigrinum)was
reduced relative to current population sizes. In general,
from 1951 to 2004, both the forest extent and adjacency
increased dramatically (Fig. 1), which likely promoted in-
creased salamander diversity and numbers. For anuran
species, the impressive numbers of B. terrestris and R.
sphenocephala witnessed in 2003 as the wetland refilled
after drought may have occurred under similar conditions
prior to 1951, although no earlier records are available.
Nonetheless, it is reasonable to assume that many of the
amphibian species, particularly the forest-dependent sala-
manders, that use Ellenton Bay as a breeding site would
not have persisted if the wetland had remained embedded
in an agricultural landscape without forested peripheral
habitats.
Our findings highlight the key role of small, isolated
wetlands in amphibian productivity and in maintaining
community dynamics by coupling aquatic habitats with
adjacent terrestrial habitats via transfer of biomass and
energy. Although the observed numbers and biomasses
of amphibians we report are very high, we suspect these
results are typical responses to postdrought conditions,
rather than a one-time occurrence. Our findings offer
hope that even moderate conservation efforts that pro-
tect wetlands and allow surrounding terrestrial habitat
to recover from prior disturbance can promote a diverse
amphibian community. Our results also suggest that cur-
rent U.S. wetland regulations that do not offer protection
to isolated wetlands will jeopardize conservation efforts
to preserve the contribution of amphibian biodiversity to
ecosystem productivity on a landscape level.
Acknowledgments
We thank R. D. Semlitsch, P. Mason, E. E. Clark, and G. J.
Graeter for providing comments on the manuscript and
the following individuals for field assistance: M. Aresco,
V. Burke, T. Carroll, E. Clark, J. Clark, K. Clark, J. Cole, L.
Cook, H. Dessauer, D. DiIullo, G. Edwards, S. Fedewa, B.
Fokidis, M. Gibbons, K. Grayson, V. Guy, W. Guthrie, I. Ha-
gen, P. Hill, W. Johnson, B. Lawrence, P. Mason, M. Mills,
T. Mills, S. Murray, B. Nelson, A. Newman, M. O’Neill, J.
Perry, A. Pickens, M. Pilgrim, S. Poppy, S. Reichbach, T.
Robison, L. Ruyle, M. Salmi, R. Semlitsch, M. Slinkard, C.
Thawley, R. Thomasson, J. Van Dyke, J. Williams, and A.
Young. We thank P. Mason, T. Mills, and S. Poppy for drift-
fence construction and repair. Research and manuscript
Conservation Biology
Volume 20, No. 5, October 2006
1464 Amphibian Biomass and Abundance Gibbons et al.
preparation were aided by the Environmental Remedia-
tion Sciences Division of the Office of Biological and Envi-
ronmental Research, U.S. Department of Energy through
Financial Assistance Award no. DE-FC09–96SR18546 to
the University of Georgia Research Foundation. Fund-
ing for M.E.D. was partially provided by National Sci-
ence Foundation grant DUE-9980743 and The Duke En-
ergy Foundation. D.A.C. was supported by a Board of Re-
gents Superior Graduate Fellowship from the University
of New Orleans. The procedures used in this study were
approved by the University of Georgia animal care and
use committee (A2003–10024, “Reptile and Amphibian
Research–General Field Studies”) and the South Carolina
Department of Natural Resources (collection permits 56–
2003, 07–2004).
Literature Cited
Barker, W. T., and G. W. Fulton. 1979. Analysis of wetland vegetation on
selected areas in southwestern North Dakota. North Dakota Regional
Environmental Assessment program report 79–15. North Dakota
State University, Fargo.
Bonin, J., J. I. DesGranges, J. Rodrigue, and M. Ouellet. 1997. Anuran
species richness in agricultural landscapes of Quebec: foreseeing
long-term results of road call surveys. Pages 141–149 in D. M. Green,
editor. Amphibians in decline: Canadian studies of a global problem.
Society for the Study of Amphibians and Reptiles, St. Louis, Missouri.
Brinson, M. M., A. E. Lugo, and S. Brown. 1981. Primary productivity, de-
composition, and consumer activity in freshwater wetlands. Annual
Review of Ecology and Systematics 12:123–161.
Burton, T. M., and G. E. Likens. 1975. Salamander populations and
biomass in the Hubbard Brook Experimental Forest, New Hamp-
shire. Ecology 56:1068–1080.
Busbee, W. S., W. H. Conner, D. M. Allen, and J. D. Lanham. 2003. Com-
position and above ground productivity of three seasonally flooded
depressional forested wetlands in coastal South Carolina. Southeast-
ern Naturalist 2:335–346.
Congdon, J. D.,J. L. Greene, and J. W. Gibbons. 1986. Biomass of freshwa-
ter turtles: a geographic comparison. American Midland Naturalist
115:165–173.
Dahl, T. E. 1990. Wetlands losses in the United States: 1780s to 1980s.
U.S. Fish and Wildlife Service, Washington, D.C.
Dahl, T. E. 2000. Status and trends of wetlands in the conterminous
United States 1986–1997. U.S. Fish and Wildlife Service, Washington,
D.C.
Davis, C. E., and L. L. Janecek. 1997. Area no. 1: field 3–412/Ellenton
Bay. Department of Energy Research Set-Aside Areas of the Savannah
River Site. SRS National Environmental Research Park Program, SRO-
NERP-25. U.S. Department of Energy, Aiken, South Carolina.
Deutschman, M. R., and J. J. Peterka. 1988. Secondary production of
tiger salamanders (Ambystoma tigrinum) in three North Dakota
prairie lakes. Canadian Journal of Fisheries and Aquatic Sciences
45:691–697.
Dodd, C. K., Jr. 1991. Drift fence-associated sampling bias of amphib-
ians at a Florida sandhills temporary pond. Journal of Herpetology
25:296–301.
Downing, D. M., C. Winer, and L. D. Wood. 2003. Navigating through
Clean Water Act jurisdiction: a legal review. Wetlands 23:475–493.
Gibbons, J. W. 1990. Chapter 1. The slider turtle. Pages 3–18 in J. W. Gib-
bons, editor.Life histor y and ecology of the slider turtle. Smithsonian
Institution Press, Washington, D.C.
Gibbons, J. W. 2003. Terrestrial habitat: a vital component for herpeto-
fauna of isolated wetlands. Wetlands 23:630–635.
Gibbons, J. W., and R. D. Semlitsch. 1981. Terrestrial drift fences with pit-
fall traps: an effective technique for quantitative sampling of animal
populations. Brimleyana 7:1–16.
Gibbons, J. W., and R. D. Semlitsch. 1991. Guide to the reptiles and
amphibians of the Savannah River Site. University of Georgia Press,
Athens.
Gibbons, J. W., J. L. Greene, and J. D. Congdon. 1983. Drought-related
responses of aquatic turtle populations. Journal of Herpetology
17:242–246.
Gibbs, J. P. 1993. Importance of small wetlands for the persistence of
local populations of wetland-associated animals. Wetlands 13:25–31.
Gibbs, J. P. 1998. Distribution of woodland amphibians along a forest
fragmentation gradient. Landscape Ecology 13:263–268.
Godley, J. S. 1980. Foraging ecology of the striped swamp snake, Regina
alleni, in southern Florida. Ecological Monographs 50:411–436.
Guerry, A. D., and M. L. Hunter Jr. 2002. Amphibian distributions in a
landscape of forests and agriculture: an examination of landscape
composition and configuration. Conservation Biology 16:745–754.
Guthery, F. S., and F. C. Bryant. 1982. Status of playas in the Southern
Great Plains. Wildlife Society Bulletin 10:309–317.
Hecnar, S. J., and R. T. M’Closkey. 1996. Regional dynamics and the
status of amphibians. Ecology 77:2091–2097.
Hecnar, S. J. 1997. Amphibian pond communities in southwestern On-
tario. Pages 1–15 in D.M. Green, editor. Amphibians in decline: Cana-
dian studies of a global problem. Society for the Study of Amphibians
and Reptiles, St. Louis, Missouri.
Himes, J. G. 2000. Burrowing ecology of the rare and elusive Louisiana
pine snake, Pituophis ruthveni (Serpentes: Colubridae). Amphibia-
Reptilia 22:91–101.
Hooper, N. M., and G. G. C. Robinson. 1976. Primary production of epi-
phytic algae in a marsh pond. Canadian Journal of Botany 54:2810–
2815.
Iverson, J. B. 1982. Biomass in turtle populations: a neglected subject.
Oecologia 59:69–76.
Knutson, M. G., J. R. Sauer, D. A. Olsen, M. J. Mossman, L. M. Heme-
sath, and M. J. Lannoo. 1999. Effects of landscape composition and
wetland fragmentation on frog and toad abundance and species rich-
ness in Iowa and Wisconsin, U.S.A. Conservation Biology 13:1437–
1446.
Lawler, G. H., L. A. Sunde, and J. Whitaker. 1974. Trout production in
prairie ponds. Journal of the Fisheries Research Board of Canada
31:929–936.
Leeper, D. A., and B. E. Taylor. 1998. Abundance, biomass and produc-
tion of aquatic invertebrates in Rainbow Bay, a temporary wetland
in South Carolina, USA. Archiv f¨ur Hydrobiologie 143:335–362.
Neckles, H. A. 1984. Plant and macroinvertebrate responses to water
regime in a whitetop marsh. M.S. thesis. University Minnesota, Min-
neapolis.
Patterson, K. K. 1978. Life history aspects of paedogenic populations of
the mole salamander, Ambystoma talpoideum. Copeia 1978:649–
655.
Petranka, J. W., and S. M. Murray. 2001. Effectiveness of removal sam-
pling for determining salamander density and biomass: a case study
in an Appalachian streamside community. Journal of Herpetology
35:36–44.
Plante, C., and J. A. Downing. 1989. Production of freshwater inverte-
brate populations in lakes. Canadian Journal of Fisheries and Aquatic
Sciences 46:1489–1498.
Porej, D., M. Micacchion, and T. E. Hetherington. 2004. Core terres-
trial habitat for conservation of local populations of salamanders
and wood frogs in agricultural landscapes. Biological Conservation
120:399–409.
Richardson, C. J. 1983. Pocosins: vanishing wastelands or valuable wet-
lands? BioScience 33:626–633.
Seigel, R. A., J. W. Gibbons, and T. K. Lynch. 1995. Temporal changes in
reptile populations: effects of a severe drought on aquatic snakes.
Herpetologica 51:424–434.
Conservation Biology
Volume 20, No. 5, October 2006
Gibbons et al. Amphibian Biomass and Abundance 1465
Semlitsch, R. D. 2000. Principles for management of aquatic-breeding
amphibians. Journal of Wildlife Management 64:615–631.
Semlitsch, R. D., and J. R. Bodie. 1998. Are small, isolated wetlands
expendable? Conservation Biology 12:1129–1133.
Semlitsch, R. D., R. N. Harris, and H. M. Wilbur. 1990. Paedomorphosis
in Ambystoma talpoideum: maintenance of population variation
and alternative life-history pathways. Evolution 44:1604–1613.
Semlitsch, R. D., D. E. Scott, J. H. K. Pechmann, and J. W. Gibbons. 1996.
Structure and dynamics of an amphibian community: evidence from
a 16-year study of a natural pond. Pages 217–248 in M. L. Cody and
J. Smallwood, editor. Long-term studies of vertebrate communities.
Academic Press, San Diego, California.
Sharitz, R. R. 2003. Carolina bay wetlands: Unique habitats of the south-
eastern United States. Wetlands 23:550–562.
Sharitz, R. R., and J. W. Gibbons. 1982. The ecology of southeast-
ern shrub bogs (pocosins) and Carolina bays. Community profile
FWS/OBS 82/04. U. S. Fish and Wildlife Service, Division of Biologi-
cal Services, Washington, D.C.
Shearer, C. R. H., and N. B. Frazer. 1997. The national environmental
research park: a new model for federal land use. American Bar Asso-
ciation’s Natural Resources and the Environment 12:46–51.
Shine, R. 1986. Ecology of a low-energy specialist: food habits and re-
productive biology of the Arafura filesnake (Acrochordidae). Copeia
1986:424–437.
Solid Waste Agency of Northern Cook County v. U.S. Army Corps of
Engineers. 2000. 191 F.3d 845 ( 7th Cir).
Stokstad, E. 2004. Global survey documents puzzling decline of amphib-
ians. Science 306:391.
Stuart, S. N., J. S. Chanson, N. A. Cox, B. E. Young, A. S. L. Rodrigues, D.
L. Fischman, and R. W. Waller. 2004. Status and trends of amphibian
declines and extinctions worldwide. Science 306:1783–1786.
Taylor, B. E., and D. L. Mahoney. 1990. Zooplankton in Rainbow Bay,
a Carolina bay pond: population dynamics in a temporary habitat.
Freshwater Biology 24:597–612.
Taylor, B. E., R. A. Estes, J. H. K. Pechmann, and R. D. Semlitsch. 1988.
Trophic relations in a temporary pond: larval salamanders and their
microinvertebrate prey. Canadian Journal of Zoology 66:2191–2198.
Taylor, B. E., D. L. Mahoney, and R. A. Estes 1989. Zooplankton pro-
duction in a Carolina bay. Pages 425–435 in R. R. Sharitz and J.
W. Gibbons, editors. Freshwater wetlands and wildlife: perspectives
on natural, managed, and degraded ecosystems. U.S. Department of
Energy, Office of Scientific and Technical Information, Oak Ridge,
Tennessee.
Taylor, B. E., D. A. Leeper, M. A. McClure, and A. E. DeBiase. 1999.
Ecology of aquatic invertebrates and perspectives on conservation.
Pages 167–196 in D. P. Batzer, R. B. Rader, and S. A. Wissinger, editor.
Invertebrates in freshwater wetlands of North America: ecology and
management. John Wiley & Sons, New York.
Taylor, B. E., D. E. Scott, and J. W. Gibbons. 2006. Catastrophic repro-
ductive failure, terrestrial survival, and persistence of the marbled
salamander. Conservation Biology (in press).
Walter, V. 2004. Object-based classification of remote sensing data for
change detection. ISPRS Journal of Photogrammetry and Remote
Sensing 58:225–238
Warner, R. R., and P. Chesson. 1985. Coexistence mediated by recruit-
ment fluctuations: a field guide to the storage effect. The American
Naturalist 125:769–787.
Whigham, D. F. 1999. Ecological issues related to wetland preservation,
restoration, creation and assessment. The Science of the Total Envi-
ronment 240:31–40.
White, D. C. 1985. Lowland hardwood wetland invertebrate commu-
nity and production in Missouri. Archiv f¨ur Hydrobiologie 103:509–
533.
Willson, J. D., and M. E. Dorcas. 2003. Effects of habitat disturbance on
stream salamanders: Implications for buffer zones and watershed
management. Conservation Biology 17:763–771.
Willson, J. D., C. T. Winne, M. E. Dorcas, and J. W. Gibbons. 2006. Post-
drought responses of semi-aquatic snakes inhabiting an isolated wet-
land: insights on different strategies for persistence in a dynamic
habitat. Wetlands: in press.
Zappalorti, R. T., and M. E. Torocco. 2002. A standardized protocol for
sampling rare snakes in the New Jersey pine barrens: critical habitat
assessment, survey techniques, and trapping methods. Herpetolog-
ical Associates, Forked River, New Jersey.
Zedler, J., et al. 2001. Compensating for wetland losses under the Clean
Water Act. National Academy Press, Washington, D.C.
Conservation Biology
Volume 20, No. 5, October 2006
