Site by Bennett Web & Design Co.
2007 SOUTHEASTERN NATURALIST 6(1):83–96
Declines in Ravine-inhabiting Dusky Salamanders of the
Southeastern US Coastal Plain
D. Bruce Means1,2,* and Joseph Travis2
Abstract - Gully-eroded and steephead valleys on Eglin Air Force Base in the Florida
panhandle were sampled for the abundance of four species of ravine-inhabiting,
plethodontid salamanders in two separate periods, 25 years apart. In this interval,
Desmognathus auriculatus (Southern Dusky Salamander) appears to have gone extinct
and the abundance of D. cf. conanti (Spotted Dusky Salamanders) has decreased by
about 68%. There was no change in the average abundance of Eurycea cirrigera (Twolined
Salamander). Pseudotriton ruber (Red Salamanders) declined in ravines from
which larger populations of D. auriculatus disappeared, but increased in ravines from
which smaller populations of D. auriculatus had disappeared. There was a slight
increase in the average abundance of P. ruber in ravines that were inhabited by D. cf.
conanti, but those changes in P. ruber abundance were unrelated to the changes in the
abundance of D. cf. conanti. Declines in populations of D. auriculatus were also noted
in Louisiana and Georgia; evidence suggests that all of these declines began in the mid-
1970s. There are several potential causes of the regional declines, but no single
explanation appears sufficient to explain declines in all populations. Feral pig rooting
eliminates the larval seepage habitat of desmognathine salamanders and may be partly
responsible for the declines on Eglin Air Force Base.
Extinction and precipitous population declines in amphibians have been
reported for many species the world over, even from relatively pristine
habitats (Houlahan et al. 2000, Stuart et al. 2004). These examples primarily
involve the anurans, frogs and toads. Disappearances of individual species
are the most well-known examples (e.g., Rheobatrachus silus Liem [Gastricbrooding
Frog; Ingram and McDonald 1993], and Bufo periglenes Savage
[Golden Toad; Crump et al. 1992]), but in some locations, entire faunas of
unrelated frog species have declined catastrophically (Laurance et al. 1996,
Lips 1998). Reports of declining salamanders have been fewer (Blaustein et
al. 1994b, Dodd 1997), less convincing (Pechmann et al. 1991), or restricted
to a small portion of the species’ range (Dodd 1997, Means et al. 1996).
The largest salamander family, Plethodontidae, has not been implicated
in precipitous declines (Hairston and Wiley 1993). Declines in plethodontid
populations would be especially important to an ecosystem because these
salamanders are often extremely abundant and, in many locations, are an
important source of high-quality energy for a variety of predators (Burton
and Likens 1975, Petranka and Murray 2001).
1Coastal Plains Institute and Land Conservancy, 1313 Milton Street, Tallahassee, FL
32303. 2Department of Biological Science, Florida State University, Tallahassee, FL
32306. *Corresponding author - firstname.lastname@example.org.
84 Southeastern Naturalist Vol. 6, No. 1
Like many streamside-inhabiting plethodontids, Desmognathus
auriculatus Holbrook (Southern Dusky Salamander) has been reported as an
abundant inhabitant of streams in the southeastern US Coastal Plain (Means
1974, 1975, 2000). While some population declines were noticed in the mid-
1970s (Means cited in Dodd 1997, 1998), these were considered to be within
the range of normal population fluctuations. A recent report by Dodd (1998)
that the species had disappeared from one locality in peninsular Florida,
however, suggested that other populations of the species might have declined.
If so, this would be the first example of a general decline in any
plethodontid and, if part of a general species-wide or area-wide pattern,
could have important consequences for the ecosystem processes in their
stream and ravine habitats.
In this paper, we use data from surveys of salamander abundance in
streams in the western Florida panhandle in two periods, twenty-five
years apart, to assess whether there is a general decline in the abundance
of D. auriculatus and the abundances of three other plethodontid salamanders
that occur in the same (Eurycea cirrigera Green [Two-lined
Salamander], Pseudotriton ruber Latreille [Red Salamander]) or similar
(D. cf. conanti Rossman [Spotted Dusky Salamander]) habitats. We report
what appear to be precipitous declines and perhaps extinctions of
many populations of D. auriculatus, declines in the populations of D. cf.
conanti, and the effects of those declines on the abundance of the
syntopic species P. ruber and E. cirrigera.
Methods and Materials
Dusky salamanders (subfamily Desmognathinae of the lungless family
Plethodontidae) are found widely distributed in the US east of the Mississippi
River, with outliers in eastern Texas, Oklahoma, and Arkansas (Conant
and Collins 1998, Petranka 1998). Many occur in or adjacent to mountain
stream habitats of the Appalachian Mountains or other interior highlands,
with the exception of one high-elevation terrestrial species and one burrower
(Petranka 1998). A number of species live in the low-elevation Coastal
Plain, a band of land skirting the southeastern corner of the US from New
Jersey to Texas. There, many populations of dusky salamanders, along with
other streamside-dwelling plethodontids, live in ravine habitats that approximate
the mountain stream habitat of the interior highlands (Means 2000).
Cool, humid ravines are critical habitats for Coastal Plain plethodontid
salamanders because ravines offer refuge from the intense mid-summer heat
and desiccation (Means 2000). The Florida panhandle is better endowed
with ravines that provide optimum plethodontid habitat close to the seacoast
than any other part of the extensive Coastal Plain from Virginia to eastern
Texas, because the area uniquely possesses a special type of ravine called
“steephead” (Means 1975, 1991, 2000). Steepheads are formed by spring
sapping of surficial groundwater aquifers in deep, porous deposits of sand.
2007 D.B. Means and J. Travis 85
Whereas classic gully-eroded ravines are usually dry in their first-order
reaches (Strahler 1964) unless it has recently rained, water flows permanently
all along steepheads where spring water emerges from the toe of steep
slopes. This water is characterized by relatively constant temperatures yearround
(19–22 C) and constant chemical composition. The perennial seeps
associated with steepheads, therefore, are ideal habitats for mountain
stream-loving salamanders (Means 2000).
Steepheads and ravines across the Florida panhandle are populated by
Red Salamanders, Two-lined Salamanders, and one of three species of
Desmognathus (Means 1975, 2000). The endemic species D. apalachicolae
Means and Karlin occupies ravines in the Apalachicola, Ochlockonee, and
lower Chattahoochee drainage basins (Means and Karlin 1989). Habitats
further west are occupied by a Desmognathus that appears to be an
undescribed species (D.B. Means, unpubl. data) that we refer to here as the
D. cf. conanti. Additionally, Means (1975) discovered that in the western
part of the Florida panhandle, several small drainages emptying independently
into the Gulf of Mexico are occupied by D. auriculatus, instead of D.
cf. conanti. Although D. auriculatus is more generally found in swampy
habitats in other areas of Florida, the other desmognathines are not. Population
genetic data is consistent with extremely low to negligible rates of
movement of D. apalachicolae and D. cf. conanti among ravines (Blouin
1986). In this paper, we present data taken in the streams that, up through the
mid-1970s, were occupied either by the D. auriculatus or D. cf. conanti.
Study area and data
The principal study area is a group of 129 deep, shaded, cool steepheads
and ravines on Eglin AFB in Santa Rosa, Okaloosa, and Walton counties, FL
(Fig. 1). We also examined data from other localities on Eglin AFB and
elsewhere in the Florida panhandle. Salamanders were collected by D.B.
Means on all visits to sites between 1969 and 1998 (some sites were sampled
more than once), with one or more other people assisting 22% of the time.
Salamanders were collected in a standard fashion: suitable-appearing microhabitats
were investigated by crawling on hands and knees and scraping
decomposing litter from the substrate using either the hands or a potato rake.
Sometimes we used the side of the boot to scrape back the top few centimeters
of decomposing litter off the top of small mucky depressions.
The data in this paper were apportioned over two time periods about 25
years apart, 1969–1975 and 1 October 1997–30 September 1998. In the
1970s, all specimens were collected and preserved. In the 1990s, all salamanders
were counted, but only a few larvae and one or two metamorphosed
specimens were collected and preserved.
Field surveys on Eglin AFB were grouped in the following manner: 1) all
localities in which D. auriculatus had been collected in the 1970s; 2) all localities
in which D. cf. conanti had been collected in the 1970s; 3) a new set of
37 localities in which D. auriculatus was endemic; and 4) a new set of 40
localities in which D. cf. conanti was endemic. Sites of the latter two groups
86 Southeastern Naturalist Vol. 6, No. 1
Figure 1. Distribution of steepheads and ravines in which (a) Desmognathus
auriculatus and (b) D. cf. conanti are endemic on Eglin Air Force Base and vicinity,
Santa Rosa, Okaloosa, and Walton counties, FL.
had not been visited previously by the authors. We collected for a minimum of
one hour at all the localities, except in those few that had been altered by
impoundment, sedimentation, or some other gross mechanical disturbance
that left the site with very little suitable habitat for plethodontids. Most of the
2007 D.B. Means and J. Travis 87
Eglin AFB ravines have remained in the same, relatively pristine ecological
condition over the 25 years of this study. No logging of the slope hardwood
forests ever took place in the steephead habitats, and the surrounding sandhill
uplands have remained undeveloped. The hydrology of steephead streams is
driven by the seepage of large, surficial aquifers so that water flow is
relatively continuous and nonfluctuating, even during the most severe
droughts (Means 1991, 2000; Wolfe et al. 1988).
Statistical analyses were performed on catch per unit effort (CPUE): in
this case, the number of animals captured per hour of searching (hereafter
denoted as rate of capture). Analyses of CPUE perform best when there is
little variation among samples in the search time. For these data, the coefficients
of variation in number caught (0.76 to 1.48) exceeded those in search
time (0.32 to 0.71). The difference is more marked in the matched comparisons
of catch and effort for each individual ravine. We assume that detection
probabilities did not change over time.
The two species of Desmognathus have declined dramatically since the
1970s, and D. auriculatus is effectively extinct in these ravines (Table 1). In
the 1970s, the average rate of capture of D. auriculatus in 26 ravines was
8.65 salamanders per hour. No salamanders of this species were seen in the
1990s, neither in the same ravines sampled in the 1970s nor in 37 novel
ravines. The change in capture rate in the 26 ravines sampled twice is
significant (matched pairs t = 7.18, P < 0.001).
In the 1970s, D. cf. conanti was more abundant than D. auriculatus,
with an average rate of capture in 26 ravines of 13.56 salamanders per
hour (Table 1). This rate was lower in 22 of the same ravines in the
1990s, an average decline of about 68%. The difference in rate of capture
between visits to the same ravines is significant (matched pairs t = 4.16, P
< 0.001). Prior sampling cannot explain this change; there is no significant
difference in rate of capture between resampled ravines and ravines
surveyed for the first time in the 1990s (average rate of 4.40 salamanders
per hour, t = 0.14, NS).
Table 1. Average rate of capture plus one standard error of four salamander species in ravines
designated by the desmognathine species that was present in the 1970s.
D. auriculatus D. cf. conanti
Species 1970s 1990s old 1990s novel 1970s 1990s old 1990s novel
D. auriculatus 8.65 (1.29) 0 0 - - -
D. cf. conanti - - - 13.56 (2.12) 4.66 (1.12) 4.40 (1.05)
P. ruber 4.95 (1.17) 5.80 (1.12) 5.83 (0.82) 2.83 (0.57) 3.61 (0.77) 3.06 (0.55)
E. cirrigera 2.59 (0.59) 2.99 (0.72) 4.60 (0.67) 3.63 (1.57) 3.67 (0.74) 2.85 (0.44)
88 Southeastern Naturalist Vol. 6, No. 1
Data for P. ruber exhibit three interesting patterns. First, in both sampling
periods and for novel or resampled ravines, P. ruber was more abundant in D.
auriculatus ravines than in D. cf. conanti ravines (Table 1). Across sampling
periods, the rates of capture of P. ruber in D. auriculatus ravines averaged
about 5.3 salamanders per hour, but only about 3.2 salamanders per hour in D.
cf. conanti ravines. The difference between ravine types was significant (F =
7.74, df = 1, 168, P = 0.006).
Second, the overall abundance of P. ruber has not changed across time.
The rates of capture of P. ruber increased by 17–27% across the decades, but
the large variation among ravines precludes this difference from being
Third, although on average there was no change in P. ruber abundance, the
changes in a subset of ravines showed a strong pattern (Fig. 2). In ravines that
had been inhabited by D. auriculatus, the change in P. ruber abundance was
positively correlated with the change in D. auriculatus abundance (r = 0.46, P less than
0.05). Ravines that suffered greater declines in D. auriculatus abundance also
had declines in P. ruber, whereas ravines that lost small populations of D.
auriculatus actually exhibited increases in P. ruber abundance. In ravines that
were inhabited by D. cf. conanti, there was no relationship between the change
in P. ruber abundance and the change in the abundance of D. cf. conanti.
The rates of capture of E. cirrigera were comparable between D.
auriculatus and D. cf. conanti ravines (3.4 and 3.5 respectively), but they
Figure 2. The relationship of the change in rate of capture of P. ruber between the
1970s and 1990s to the change in the rate of capture of D. auriculatus in the same
ravines across the same interval.
2007 D.B. Means and J. Travis 89
exhibited no consistent change across sampling periods or ravine types
(Table 1). Neither the magnitude nor the direction of change in the rate of
capture of E. cirrigera in a ravine was related to either the type of ravine (D.
auriculatus or D. cf. conanti) or the magnitude of change in the
desmognathine population in the ravine.
It is unlikely that these results are artifacts of sampling. The percentage
decrease in average rates of capture of the two Desmognathus species
between decades far exceeds the percentage decrease in average sampling
durations between the 1970s (1.60 hours, s.e. = 0.20) and 1990s (1.31 hours,
s.e. = 0.19). Moreover, rates of capture of the other species were either
higher (P. ruber) or comparable (E. cirrigera) in the 1990s to the rates in the
1970s, which should not be the case if our results were driven entirely by the
slightly shorter sampling durations in the 1990s. While our method did not
control for a variety of factors that can affect detection probabilities (e.g.,
recent rainfall history, air temperature; see MacKenzie et al. 2002), variation
in these factors should have increased the variance among ravines within
each period, which makes the difference in rates of capture between periods
more notable and the statistical tests of these effects conservative.
The decline of D. auriculatus
Once the most abundant salamander in the steepheads and floodplain
swamps it occupied, D. auriculatus now appears to be entirely absent from
the 185,600-ha Eglin Air Force Base. It is possible, at least in theory, that by
sampling only twice, we have inadvertently sampled high and low points of
normal population fluctuations. However, for this to be the case, all 129
populations of the two desmognathine species would have to be fluctuating
synchronously, which seems unlikely. Even if this were the case and all
desmognathines were in the same numerical trough, we would have expected
to find at least a few individual D. auriculatus in the nearly 80 hours
spent searching 63 ravines.
Desmognathus auriculatus is also missing from several other localities in
Florida where it had been abundant in the early 1970s (based upon collection
records in the Florida State and Louisiana State museums and the Coastal
Plains Institute). D. auriculatus was abundant in the Ochlockonee River
floodplain (below Old Bainbridge Road in Leon County), but have not been
seen since 1971. In the Telogia Creek floodplain (Liberty County), no D.
auriculatus have been seen since 1974. The last Southern Dusky Salamander
in Deep Springs Canyon, a large steephead of Econfina Creek (Bay County),
was observed in 1976, in spite of many efforts to find it there in the 1990s
(D.B. Means, pers. observ.).
Perhaps the most striking decline, for historical reasons, has occurred at
Silver Glen Springs, FL (Marion County). Neill (1951) brought formal attention
to this population by naming it D. f. carri, which was subsequently placed
in the synonomy of D. auriculatus (Means 1974, Rossman 1959). Neill’s
90 Southeastern Naturalist Vol. 6, No. 1
(1951) type series contained a collection of 22 juveniles and adults taken on 12
November 1950 and another of 18 juveniles and adults collected on 29 January
1951. On 25 November 1958, Rossman (1959) found D. auriculatus abundant
at this site, collecting 36 specimens, and Christman’s (1970) report was based
on 91 specimens he collected from three sites including Silver Glen Springs. D.
Rossman ( LSU Museum of Natural History, Baton Rouge, LA, pers. comm.,
1999) said that Silver Glen Springs was the most robust population of D.
auriculatus he had seen at that time. On 16 February 1972, S.P. Christman and
D.B. Means found 9 specimens in one person-hour of searching in the vicinity
of the spring boils, and the species was common in the surrounding hardwood
bottomland swamps (n = 10; 1.5 person-hours). On three collecting trips
between June 1994 and March 1995 to collect D. auriculatus at Silver Glen
Springs, R. Franz and K. Dodd were unsuccessful (Dodd 1998), but Christman
found two or three individuals in November 1995 (Dodd 1998).
The four habitats discussed above were high quality, extensive habitats that
previously supported large populations of D. auriculatus, but which produced
few or no specimens after about 1975. In fact, there are only two localities in the
Florida panhandle where D. auriculatus has been found recently: Monkey
Creek in the Bradwell Bay Wilderness Area on the Apalachicola National
Forest in Wakulla County and a large cypress swamp on the Apalachicola
National Forest in southern Liberty County, about 18 km west of the Monkey
Creek site. Both sites are acidic, blackwater swamps and are very different
habitats in comparison with the rheophilic ravines.
Desmognathus auriculatus has also declined or been extirpated in other
states. At a site on Anderson Branch of Hunters Creek, GA (Irwin County), one
larva and 26 juveniles and adults were collected on 15 December 1971 and 25
January 1972 in 6 person-hours of effort. The site was revisited 14 years later in
1986, and no D. auriculatus were found in 3 hours of intensive collecting, while
on 18 March 1995, with 4 hours of vigorous searching, one small juvenile was
found. Of 2789 D. auriculatus specimens in southern Louisiana museum
collections from the Florida parishes, only 14 (0.5%) were collected after 1975.
On the basis of localities, D. auriculatus was collected from only 2 of 75
historic localities since 1980. On the other hand, 21% of the D. cf. conanti in
collections were taken after 1975, and the species has been collected in 11 of 71
localities since 1980 (Boundy 2005; J. Boundy, Louisiana Department of
Wildlife and Fisheries, Baton Rouge, LA, pers. comm.).
Based on the accumulated evidence, a widespread decline or extirpation
of D. auriculatus has taken place in many Coastal Plain localities. In the
ravines on Eglin AFB, D. cf. conanti has also declined. It is important to
note that all of these declines may have begun at approximately the same
time, in the mid-1970s. A similar conclusion was reached for the timing of
the extinction of D. auriculatus in Devil’s Millhopper in central Florida
(Dodd 1999). We do not have comparable data on abundance of D.
apalachicolae or the rarer D. monticola Dunn, and there is no evidence
from our field work, or that of other herpetologists in the area, that these
species have undergone any comparable decline.
2007 D.B. Means and J. Travis 91
Changes in P. ruber
The abundance of P. ruber in ravines inhabited by D. cf. conanti showed
no significant change between the 1970s and 1990s, but their abundance in
D. auriculatus ravines changed in proportion to the magnitude of the D.
auriculatus decline. Ravines that had held larger populations of D. auriculatus
showed declines in their P. ruber populations and ravines that had
held smaller desmognathine populations showed increases. One hypothesis
for this pattern is that the loss of Southern Dusky Salamanders represented a
significant loss of prey for P. ruber, which is a well-known predator of other
salamanders (Gustafson 1993, Petranka 1998). The failure to see a similar
pattern in the D. cf. conanti ravines could be ascribed to the possibility that
D. auriculatus was a more important resource for P. ruber than is D. cf.
conanti, a possibility bolstered by the observation that, in the 1970s, average
P. ruber abundance in D. auriculatus ravines was 75% higher than their
average abundance in D. cf. conanti ravines. Of course, a second, equally
viable hypothesis is that the changes in P. ruber abundance were driven by
factors other than changes in the desmognathine abundance; those factors
could be the same as those that affected the desmognathines or they could be
Potential causes of the declines of desmognathine populations
Any explanation for the widespread declines and extinctions in D.
auriculatus and D. cf. conanti must account for the absence of any negative
effect on the syntopic P. ruber and E. cirrigera or the abundance of the other
members of the genus Desmognathus in north Florida. The major hypotheses
include one or more of the effects posited to cause amphibian declines in
other regions (overcollecting, acid rain, ultraviolet-B radiation, toxic substances,
disease), and the indirect effects of Sus scrofa Linnaeus (feral pigs)
on habitat structure.
It is possible that different factors have combined to affect different
populations and that there is no single cause for these declines. The simplest
explanation is that D.B. Means over-collected every population that he
sampled in the 1970s. This hypothesis is unlikely for three reasons. First,
quantitative studies in other streamside communities indicate that a far
greater collecting effort would be necessary to create a long-term loss of
animals (Hairston 1986, Petranka and Murray 2001). Second, only the
desmognathines declined, not the hemidactylines, which were collected in
the same samples. Third, the hypothesis is inconsistent with the fact that the
densities of the desmognathines were just as low in ravines that had not been
sampled in the 1970s as in those that had. In addition, the population in
Monkey Creek within the Bradwell Bay Wilderness Area seems to be thriving
despite periodic collections by D.B. Means from this population.
Three causes of other amphibian declines—acid rain, ultraviolet-B (UV-B)
radiation, and toxic substances—are unlikely to be important in these cases.
For one reason, the effects would have to be confined to two species of
Desmognathus and not to other, syntopic plethodontids. In addition, the
92 Southeastern Naturalist Vol. 6, No. 1
habitats used by these species seem unlikely to be affected by these factors.
Florida receives fairly high levels of acid rain (Brezonik et al. 1980), but D.
auriculatus inhabits some of the most naturally acid wetland habitats in the
Coastal Plain (Means 1999). In fact, as already mentioned, it is in acid swamps
that known populations still exist. One might expect acid rain to affect D. cf.
conanti, which exclusively occupies ravines in which aquifer water or rainfall
is circumneutral to slightly acid, yet this species remains extant, albeit in lower
abundances. While ambient ultraviolet-B (UV-B) radiation causes significant
embryonic mortality in some amphibian species (Anzalone et al. 1998;
Blaustein et al. 1994a, 1995), the more susceptible populations studied to date
occur at high elevations (1190–2000 m) and breed in shallow lakes and other
wetlands that are open to full sunlight. The habitats of the plethodontid
salamanders reported here are much lower in elevation (0–200 m) and densely
forested. In addition, the plethodontids live in microhabitats that are naturally
protected from UV-B such as in decomposing leaf litter and water, or under
logs, rocks, and the soil surface beneath moss or fern ground cover. While toxic
substances such as heavy metals, herbicides, and pesticides cannot be discounted
(Diana and Beasley 1998), their effects must be very selective among
species in order to account for the patterns documented here.
Disease caused by a microbial pathogen is a plausible cause of the
declines noted here (Faeh et al. 1998). Differential susceptibility to a pathogen
could be correlated with phylogenetic affinity, which would account for
the selective nature of the declines and the differential severity between D.
auriculatus and D. cf. conanti. The persistence of populations in and around
the Monkey Creek area might be explained by a simple failure of the
pathogen to invade those populations, even though it would have invaded all
of the surrounding populations.
The effect of feral pigs on habitat structure is a prime candidate for some
of the declines. Feral pigs eat and uproot plants (Hardin 1994, Lipscomb
1989), eat animals (Douglass and Winegarner 1977, Wood and Roark 1980),
modify the soil by mixing organic and mineral layers (Ebenhard 1988), and
mechanically disrupt microtopographic relief of seepage wetlands, thereby
altering microhydrology (Layne 1997; D.B. Means, pers. observ.; Randall et
al. 1997). The first evidence of feral pig rooting in an Eglin AFB steephead
was not noticed until 26 February 1975. However, during 1997–1998, severe
feral pig damage was observed in 77 of 160 (48%) steepheads and ravines,
moderate damage in 19 of 160 (12%), light damage in 3 of 160 (2%), and
little or no evidence of recent pig rooting was recorded for 61 (38%)
steepheads and ravines.
There are limited data that feral pigs eat salamanders (Singer et al. 1982,
Springer 1977), but the more likely effect of feral pig rooting is to alter the
larval habitat of the desmognathines, and perhaps thereby have a greater effect
on these species than the hemidactylines. The larvae of both D. auriculatus and
D. cf. conanti in steepheads are found on small, sandy, seepage sites with only a
thin sheet-flow (2–5 mm) of water and covered with a veneer of decomposing,
2007 D.B. Means and J. Travis 93
multicolored leaf litter. Psuedotriton ruber hatchlings are found in such
habitats for a short while until they are about the size of the largest
desmognathine larvae, at which time they move into deeper, mucky pools. The
larvae of E. cirrigera are most commonly found in flowing water where
seepage collects in small rivulets, and larger E. cirrigera larvae are also found
in mucky or wet peaty sites with P. ruber. Suitable habitat for desmognathine
larvae is confined to the upper reaches of the steephead, whereas patches of
mucky habitat for hemidactyline larva occur throughout the stream course.
Repeated pig rooting transforms the gentle seepage slopes that are habitat for
desmognathine larvae into flat, submerged beds of deep organic matter—
exactly the habitats preferred by the larger P. ruber and E. cirrigera larvae.
Besides adult salamanders hiding under leaf and twig litter on seepage slopes,
Diplocardia mississippiensis Smith (earthworms) in the soil are probably what
pigs are seeking in the seepage slopes. Once a seepage slope has been converted
to a mucky pool, further pig rooting does not change the habitat preferred by the
While it is likely that feral pigs have affected the desmognathines in Eglin
AFB ravines, the wider disappearance of D. auriculatus from diverse localities
across the species range in which there is no evidence of feral hog depredations
suggests either that a more pervasive agent, such as a pathogen, is acting or that
there is no single cause for the species decline. Those few localities (Monkey
Creek, Apalachicola National Forest) in which D. auriculatus still can be found
are highly acidic, flatwoods swamps. Possibly, the acidity of blackwater
environments precludes pathogens such as viruses, bacteira, and chytrid fungi,
all known pathogens that negatively affect amphibians (Lannoo 2005). A
range-wide survey of all known populations of D. auriculatus, plus a search for
new localities, should be mounted immediately. There should also be a search
for any candidate pathogens in this species. Discovering declines such as those
reported here, and determining their cause, highlights the importance of
establishing a nationwide biomonitoring program (Bishop and Pettit 1992,
Kim and Knutson 1986, National Research Council 1993).
For help in the field over the years, we thank Harley Means, Ryan Means, Jim
Berry, Storrs Olson, Dan Simberloff, Don Strong, Bob Godfrey, Prince Jinright,
Bobby Crawford, Clive Longden, Wilson Baker, Al Karlin, David Printiss, and Jim
Eggert. We thank David Printiss and the Florida Natural Areas Inventory for their
cooperation and information on geographical distributions of animals, plants, and
ecosystems on Eglin AFB. We thank Michael Turtora and Steve Laird for advice with
statistical treatment of the data. This study was partly funded by a grant from the
Legacy Resource Management Program of the Department of Defense for a survey of
amphibians inhabiting steepheads and ravines on Eglin Air Force Base (October 1997–
September 1998). We thank Carl Petrick for assistance with a grant and logistical
support on Eglin AFB. Field work was also supported by Tall Timbers Research
Station (1969–1983) and the Coastal Plains Institute and Land Conservancy (1984–
2002). J. Travis’ research is supported by the National Science Foundation through
94 Southeastern Naturalist Vol. 6, No. 1
award DEB 99-03925. For a critical reading of the manuscript, we thank C. Kenneth
Dodd, Jr. and Paul E. Moler.
Anzalone, C.R., L.B. Kats, and M.S. Gordon. 1998. Effects of solar UV-B radiation on
embryonic development in Hyla cadaverina, Hyla regilla, and Taricha torosa.
Conservation Biology 12(3):646–653.
Bishop, C.A., and K.E. Pettit (Eds.). 1992. Declines in Canadian amphibian populations:
Designing a national monitoring strategy. Canadian Wildlife Service, Occasional
Paper No. 76, 120 pp.
Blaustein, A.R., P.D. Hoffman, D.G. Hokit, J.M. Kiesecker, S.C. Walls, and J.B.
Hays. 1994a. UV repair and resistance to solar UV-B in amphibian eggs: A link
to population declines? Proceedings of the National Academy of Science
Blaustein, A.R., D.B. Wake, and W.P. Sousa. 1994b. Amphibian declines: Judging
stability, persistence, and susceptibility of populations to local and global extinctions.
Conservation Biology 8:60–71.
Blaustein, A.R., B. Edmond, J.M. Kiesecker, J.J. Beatty, and D.G. Hokit. 1995.
Ambient ultraviolet radiation causes mortality in salamander eggs. Ecological
Blouin, M.S. 1986. Regional and local patterns of electrophoretic variation in
Desmognathus: The importance of physical barriers to gene flow. M.Sc. Thesis.
Florida State University, Tallahassee, FL. 62 pp.
Boundy, J. 2005. Museum collections can assess population trends. Pp. 295–299, In M.
Lannoo (Ed.). Amphibian Declines: The Conservation Status of United States
Species. University of California Press, Berkeley, CA. 1025 pp.
Brezonik, P.L., E.S. Edgerton, and C.D. Hendry. 1980. Acid precipitation and sulfate
deposition in Florida. Science 208:1027–1029.
Burton, T.M., and G.E. Likens. 1975. Energy flow and nutrient cycling in salamander
populations in the Hubbard Brook experimental forest, New Hampshire. Ecology
Christman, S.P. 1970. An examination of the desmognathine salamanders from Silver
Glen Springs, Florida. Unpublished report in Herpetology Library, Florida Museum
of Natural History, Gainesville, FL. 23 pp.
Conant, R., and J.P. Collins. 1998. A Field Guide to Reptiles and Amphibians, Eastern
And Central North America. Third Edition, expanded. Houghton Mifflin Co.,
Boston, MA. 616 pp.
Crump, M.L., F.P. Hensley, and K.L. Clark. 1992. Apparent decline of the Golden
Toad: Underground or extinct? Copeia 1992:413–420.
Diana, S.G., and V.R. Beasley. 1998. Chapter 27. Amphibian toxicology. Pp. 266–277,
In M.J. Lanoo (Ed.). Status and Conservation of Midwestern Amphibians. University
of Iowa Press, Iowa City, IA. 526 pp.
Dodd, Jr., C.K. 1997. Chapter 6. Imperiled amphibians: A historical perspective. Pp.
165–200, In G.W. Benz and D.E. Collins(Eds.). Aquatic Fauna in Peril: The
Southeastern Perspective. Special Publication 1, Southeast Aquatic Research Institute,
Lenz Design and Communications, Decatur, GA. 554 pp.
Dodd, Jr., C.K. 1998. Desmognathus auriculatus at Devil’s Millhopper State Geological
Site, Alachua County, Florida. Florida Scientist 61(1):38–45.
Douglass, J.F., and C.E. Winegarner. 1977. Predation of eggs and young of the Gopher
Tortoise, Gopherus polyphemus (Reptilia, Testudines, Testudinidae), in southern
Florida. Journal of Herpetology 11:236–238.
2007 D.B. Means and J. Travis 95
Ebenhard, T. 1988. Introduced birds and mammals and their ecological effects.
Swedish Wildlife Research (Viltrevy) 13:1–107.
Faeh, S.A., D.K. Nichols, and V.R. Beasley. 1998. Chapter 26. Infectious diseases of
amphibians. Pp. 259–265, In M.J. Lannoo (Ed.). Status and Conservation of
Midwestern Amphibians. University of Iowa Press, Iowa City, IA. 526 pp.
Gustafson, M.P. 1993. Intraguild predation among larval plethodontid salamanders: A
field experiment in artificial stream pools. Oecologia 96:271–275.
Hairston, N.G. 1986. Speciespacking in Desmognathus Salamanders: Experimental
demonstration of predation and competition. American Naturalist 127
Hairston, Sr., N.G., and R.H. Wiley. 1993. No decline in salamander
(Amphibia:Caudata) populations: A twenty-year study in the southern Appalachians.
Hardin, D.E. 1994. Non-indigenous species on forestry-managed lands. Pp. 139-145, In
D.C. Schmitz and T.C. Brown (Eds.). An Assessment of Invasive Non-indigenous
Species in Florida’s Public Lands. Florida Department of Environmental Protection,
Technical Report TSS-94-100. Tallahassee, FL. 303 pp.
Houlahan, J.E., C.S. Findley, B.R. Schmidt, A.H. Meyer, and S.L. Kuzmin. 2000.
Quantitative evidence for global amphibian population declines. Nature
Ingram, G.J., and K.R. McDonald. 1993. An update on the decline of Queensland’s
frogs. Pp. 297–303, In D. Lunney and D. Ayers (Eds.). Herpetology in Australia. A
Diverse Discipline. Royal Zoological Society of New South Wales, Mosman,
Australia. 414 pp.
Kim, K.C., and L. Knutson (Eds.). 1986. Foundations for a National Biological Survey.
Association of Systematics Collections, Lawrence, KS. 215 pp.
Lannoo, M. (Ed.). 2005. Amphibian Declines: The Conservation Status of United States
Species. University of California Press, Berkeley, CA. 1025 pp.
Laurance, W.F., K.R. McDonald, and R. Speare. 1996. Epidemic disease and the
catastrophic decline of Australian rain forest frogs. Conservation Biology
Layne, J.N. 1997. Chapter 10. Nonindigenous mammals. Pp. 157–186, In D.
Simberloff, D.C. Schmitz, and T.C. Brown (Eds.). Strangers in Paradise: Impact and
Management of Nonindigenous Species in Florida. Island Press, Washington, DC.
Lips, K.R. 1998. Decline of a tropical montane amphibian fauna. Conservation Biology
Lipscomb, D.J. 1989. Impacts of feral hogs on longleaf pine regeneration. Southern
Journal of Applied Forestry 13(4):177–181.
MacKenzie, D.I., J.D. Nichols, G.B. Lachman, S. Droege, J.A. Royle, and C.A.
Langtimm. 2002. Estimating site occupancy rates when detection probabilities are
less than one. Ecology 83:2248–2255.
Means, D.B. 1974. The status of Desmognathus brimleyorum Stejneger and an analysis
of the genus Desmognathus (Amphibia: Urodela) in Florida. Bulletin of the Florida
State Museum, Biological Sciences 18:1–100.
Means, D.B. 1975. Competitive exclusion along a habitat gradient between two species
of salamanders (Desmognathus) in western Florida. Journal of Biogeography
Means, D.B. 1991. Florida’s steepheads: Unique canyonlands. Florida Wildlife
Means, D.B. 1999. Desmognathus auriculatus. Catalogue of American Amphibians
and Reptiles 681.1–681.6.
96 Southeastern Naturalist Vol. 6, No. 1
Means, D.B. 2000. Chapter 14. Southeastern US Coastal Plain habitats of the
Plethodontidae: The importance of relief, ravines, and seepage. Pp. 287–302, In
R.C. Bruce, R.J. Jaeger, and L.D. Houck (Eds.). The Biology of Plethodontid
Salamanders. Plenum Publishing Corporation, New York, NY. 485 pp.
Means, D.B., and A.A. Karlin. 1989. A new species of Desmognathus from the eastern
Gulf Coastal Plain. Herpetologica 45:37–46.
Means, D.B., J.G. Palis, and M. Baggett. 1996. Effects of slash pine silviculture on a
Florida population of the Flatwoods Salamander. Conservation Biology
National Research Council. 1993. A Biological Survey for the Nation. National
Academy Press, Washington, DC. 205 pp.
Neill, W.T. 1951. A new subspecies of dusky salamander, genus Desmognathus,
from south-central Florida. Publications of the Research Division Ross Allen’s
Reptile Institute 1(3):25–38.
Pechmann, J.H.K, D.E. Scott, R.D. Semlitch, J.P. Caldwell, L.J. Vitt, and J.W.
Gibbons. 1991. Declining amphibian populations: The problem of separating
human impacts from natural fluctuations. Science 253:892–895.
Petranka, J.W. 1998. Salamanders of the United States and Canada. Smithsonian
Institution Press, Washington, DC. 587 pp.
Petranka, J.W., and S.S. Murray. 2001. Effectiveness of removal sampling for determining
salamander density and biomass: A case study in an Appalachian streamside
community. Journal of Herpetology 35:36–44.
Randall, J.M., R.R. Lewis III, and D.R. Jensen. 1997. Chapter 12. Ecological restoration.
Pp. 205–219, In D. Simberloff, D.C. Schmitz, and T.C. Brown (Eds.).
Strangers in Paradise: Impact and Management of Nonindigenous Species in
Florida. Island Press, Washington, DC. 467 pp.
Rossman, D. 1959. Ecosystematic relationships of the salamanders Desmognathus
fuscus auriculatus Holbrook and Desmognathus fuscus carri Neill. Herpetologica
Singer, F.J., W.T. Swank, and E.E.C. Clebsch. 1982. Some ecosystem responses to
European wild boar rooting in a deciduous forest. Research/Resources Management
Report No. 54 USDI, NPS, SERO, Atlanta, GA. 31 pp.
Springer, M.D. 1977. Ecologic and economic aspects of wild hogs in Texas. Pp. 37–
46, In G.W. Wood (Ed.). Research and Management of Wild Hog Populations.
The Belle W. Baruch Forest Science Institute of Clemson University,
Georgetown, SC. 113 pp.
Strahler, A.N. 1964. Section 4-II. Geology. Part II. Quantitative geomorphology of
drainage basins and channel networks. Pp. 4-39 to 4-76, In Ven te Chow (Ed.).
Handbook of Applied Hydrology: A Compendium of Water-resources Technology.
McGraw-Hill, New York, NY. 1418 pp.
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 (5702):1783–1786.
Wolfe, S.H., J.A. Reidenauer, and D.B. Means. 1988. An ecological characterization
of the Florida panhandle. US Fish and Wildife Service Biological Report
Wood, G.W., and D.N. Roark. 1980. Food habits of feral hogs in coastal South Carolina.
Journal of Wildlife Management 44(2):506–511.