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Aquatic Fauna in Peril: The Southeastern Perspective Edited By George W. Benz And David E. Collins

Aquatic Fauna in Peril > Imperiled Amphibians: A Historical Perspective

Imperiled Amphibians: A Historical Perspective

By C. Kenneth Dodd, Jr.

Amphibians have been in and out of the news over the past few years because of the often unexplained disappearance of individual species or groups of species (Barinaga, 1990; Blaustein and Wake, 1990; Vitt et al., 1990; Wyman, 1990; Anonymous, 1991; Wake, 1991; Livermore, 1992, Blaustein et al., 1994c; Stebbins and Cohen, 1995). Amphibian declines or extinctions have been particularly apparent in the western United States (e.g., Bradford, 1991; Carey, 1993a; Fellers and Drost, 1993) and Australia (Richards et al., 1993), with scattered declines reported in Central and South America, Europe, and elsewhere (Vial and Saylor, 1993). Although much debate centers on natural population fluctuations (Pechmann et al., 1991; Blaustein, 1994; Pechmann and Wilbur, 1994) and their role in population viability (Sjogren, 1991a), there seems to be little doubt that amphibian populations are threatened by an ever expanding human population.

In this paper, I present an overview of the taxonomic diversity of the amphibians of the southeastern United States, the types of habitats used by amphibians, amphibian life history in relation to aquatic habitats, the types of studies that have been conducted on southeastern amphibians, and the status of and threats to particular species and populations. I define the southeastern United States to include an area from Virginia, West Virginia, and Kentucky, south through Florida and west to eastern Texas. As such, the region includes the contiguous southern Appalachians, southeastern Coastal Plain, Interior Highlands, and Edwards Plateau, all areas of important species richness and diversity.

Taxonomic Review

Of the estimated 4,300 to 4,500 amphibian species worldwide (Vial and Saylor, 1993; Zug, 1993; McDiarmid, 1994), 147 species have been described from the southeastern United States. In addition to the described species, a number of species await formal taxonomic description, particularly in the salamander families Plethodontidae, Proteidae, and Sirenidae (P. Moler, Florida Game and Fresh Water Fish Commission, pers. comm.), and possibly in the frog family Ranidae (R. Franz, Florida Museum of Natural History, pers. comm.). Within North America, the Southeast has the greatest amphibian species richness (Kiester, 1971). Most of the native amphibians in the Southeast (68 percent) are salamanders, with 99 described species. The amphibian species richness of each southeastern state is shown in Figure 1.

Order Caudata

Of the seven salamander families in the southeastern United States, two (Amphiumidae and Sirenidae) are endemic to the region while two additional families (Ambystomatidae and Proteidae) have their greatest species richness in the Southeast. One of the three extant cryptobranchids occurs primarily in southern streams and rivers, whereas the other species (Andrias spp.) are found in Asia. The family Plethodontidae exhibits substantial diversity in the Southeast, although its greatest species richness occurs in the mountainous Neotropics of southern Mexico and Central America. The family Salamandridae is primarily Palearctic and Oriental in distribution, although all three known species of Notophthalmus are found in the Southeast.

The following salamander genera have their centers of distribution within the Southeast: Cryptobranchus, Necturus, Amphiuma, Siren, Pseudobranchus, Leurognathus, Phaeognathus, Haideotriton, Stereochilus, and Typhlomolge. Most or all species of Notophthalmus, Desmognathus, Eurycea, Gyrinophilus, Plethodon, and Pseudotriton also occur in the Southeast, although the ranges of individual species may extend substantially northward.

Order Anura

There are no endemic families of frogs in the southeastern United States, and only two genera (Acris and Pseudacris) have centers of species richness within the Southeast. The highest diversity (19 species) of southeastern frogs occurs within the Hylidae (treefrogs) followed by the Ranidae (true frogs: 14+ species) and the Bufonidae (toads: seven species). Hypopachus and Syrrhophus barely enter the coastal plain in Texas, whereas Gastrophryne and Scaphiopus are together represented by a total of five species, three of which barely enter the Southeast. In addition to the 48 native frog species, at least four exotic species (Bufo marinus, Eleutherodactylus coqui, E. planirostris, Osteopilus septentrionalis) have established breeding populations (all in Florida).

Distribution And Habitats

Physiographic Regions and Centers of Species Diversity

Amphibians are found in all physiographic regions of the southeastern United States (Table 1). They are found from sea level to the tops of the highest Appalachian Mountains. Centers of species richness and endemism include the following: the Appalachian Mountains, particularly at higher elevations (salamanders, especially the family Plethodontidae and the genus Plethodon); the Atlantic and Gulf of Mexico coastal plains (many salamanders and frogs, especially Amphiuma, Siren, Pseudobranchus, Necturus, Haideotriton, and Pseudacris species); the Interior Highlands, including the Boston, Ouachita, and Ozark mountains (many endemic salamanders); and the Edwards Plateau (many endemic cave and spring salamanders of the genera Eurycea and Typhlomolge) (Figure 2).

Aquatic Habitats

Amphibians are found in all types of aquatic wetlands (see Hackney et al., 1992, and references therein) except those associated with the saline waters along the coast. Even there, however, some species occasionally are found in brackish habitats (Neill, 1958; Christman, 1974). Selected references on amphibian species composition of southeastern aquatic environments include the following: Moler and Franz (1988), LaClaire and Franz (1991), Dodd (1992), and Cash (1994) for temporary ponds; Adams and Lacki (1993) for road-ruts; Turner and Fowler (1981) and Lacki et al. (1992) for ponds at former mine sites; Delis (1993) and O’Neill (1995) for wetlands in pine flatwoods; Mitchell et al. (1993) for saturated forested wetlands; Harris and Vickers (1984) and Vickers et al. (1985) for cypress domes; Pearson et al. (1987) for bayheads; Wright (1932), Delzell (1979), and Hall (1994) for large swamps; Dalrymple (1988) for wet prairies; Parker (1937) and Bancroft et al. (1983) for lakes; and Southerland (1986) for streams. Much information on amphibian use of aquatic habitats is contained in state or regional books (e.g., Carr, 1940; Duellman and Schwartz, 1958; Ashton and Ashton, 1988; Dundee and Rossman, 1989; Gibbons and Semlitsch, 1991) as well as in numerous accounts of species in need of conservation (see Table 2).

Large, fully aquatic salamanders (Cryptobranchus, Necturus) typically are found in the larger rivers and streams, whereas small aquatic salamanders (Desmognathus, Leurognathus, Eurycea) frequent small streams and seeps. For these salamanders, larval development occurs within the stream and, after metamorphosis, adults live along the wet streamsides or among the gravelly substrate. Salamanders belonging to Siren, Pseudobranchus, and Amphiuma inhabit various types of vegetated ponds and mucky swamps. Newts and most Ambystoma species require temporary ponds to complete metamorphosis, and premature pond drying is an ever present threat to their development (Semlitsch, 1987; Pechmann et al., 1989; Dodd, 1993). Of course, even salamanders that do not require water to breed need moist environments to prevent desiccation.

As with salamanders, frogs use a variety of wetlands for reproduction. Most frog species have tadpoles which develop within ponds, lakes, wet prairies, or other lentic waters. Fewer species use streams, rivers, or swift flowing waters (e.g., Rana heckscheri in rivers, streams, and oxbows in addition to lentic waters). Some frogs are very habitat specific, such as Rana capito and Hyla gratiosa, which require fishless temporary ponds for reproduction. Some species, such as Bufo terrestris, breed in a wide variety of wetland habitats.

Terrestrial Habitats

Although amphibians are usually associated with water, most species spend a substantial amount of time in terrestrial habitats. Individuals of some species often can be found at great distances from the nearest breeding ponds. For example, I have funnel-trapped many small frogs and salamanders in the harsh Florida sandhills 200 to greater than 800 m (656 to more than 2,624 feet) from the nearest water (Dodd, 1996). Franz et al. (1988) recorded a gopher frog (Rana capito) at a tortoise burrow 2 km (1.25 miles) from where the frog had been previously marked. Such long distance movements probably are not unusual. Greenberg (1993) captured southern toads (Bufo terrestris), eastern narrow-mouthed toads (Gastrophryne carolinensis), and eastern spadefoot toads (Scaphiopus holbrooki) in Florida sand pine scrub between 5 and 6 km (3.1 and 3.7 miles) from the nearest known water source.

Terrestrial refugia include caves (Saugey et al., 1988; Franz et al., 1994); burrows of tortoises (Jackson and Milstrey, 1989), pocket gophers, crayfish (especially by Rana areolata and R. capito) and other invertebrates; tree roots; rock crevices; surface debris; and probably many other subterranean habitats. Treefrogs often use arboreal retreats. Selected references on the use of terrestrial habitats by amphibians that require water to breed include Gibbons and Bennett (1974), Bennett et al. (1980), Semlitsch (1981), Campbell and Christman (1982), Pearson et al. (1987), Stout et al. (1988), Scott (1991), McCoy and Mushinsky (1992), and Dodd (1996).

Aquatic Amphibian Life History

In North America, most amphibians have a biphasic life cycle consisting of an aquatic egg and larval stage, metamorphosis into a terrestrial adult, and migration back to water to breed and lay eggs. The time between metamorphosis and first breeding varies among species, although this period is usually one to four years (Duellman and Trueb, 1986). The life span of wild individuals also varies. For example, Gastrophryne carolinensis may live four or more years (Dodd, 1995a), whereas the hellbender may live greater than 25 years (Peterson et al., 1983). Generally, salamanders live longer than frogs, and larger species live longer than smaller species (Duellman and Trueb, 1986). Duellman and Trueb (1986) discussed life history variation and the factors that affect reproduction, life cycles, and other facets of amphibian biology.

There are exceptions to the "typical" amphibian life cycle. All non-hemidactyliine salamanders of the family Plethodontidae (i.e., Aneides and Plethodon species), two species of Desmognathus (D. aeneus and D. wrighti), and Phaeognathus hubrichti skip the aquatic larval stage (Table 1). Instead, eggs are laid on land in moist environments, the larval stage is passed within the egg, and the hatchling resembles a miniature adult.

Several salamanders, including all Siren spp., Pseudobranchus spp., Necturus spp., and Typhlomolge spp., some Eurycea spp., and Haideotriton wallacei and Cryptobranchus alleganiensis, are entirely aquatic and never leave the water or boggy wetlands. Eggs are deposited in vegetation, debris, or under rocks, young usually pass through a larval stage, and adults often retain larval features, such as exposed gills. Amphiuma spp. generally are aquatic, although eggs are deposited on land near water. Other species (Ambystoma talpoideum, Notophthalmus spp.) have individuals or populations that are facultative paedomorphs (that is, they become reproductively active while otherwise retaining larval phenotypes, and they never transform into adults while permanent water remains).

All native southeastern frogs, with the exception of the direct developing Syrrhophus spp., have a "typical" amphibian life cycle. Both of the introduced Eleutherodactylus spp. are direct developers with no aquatic life stage.

Studies And Literature On Imperiled Southeastern Amphibians

Prior to the second half of the 20th century, most studies of the aquatic Amphibia of the southeastern United States focused on general distribution patterns, morphological systematics, and life history field observations. The literature stemming from those studies has been summarized in several major monographs and field guides (e.g., for Alabama see Mount, 1975; for Florida see Carr, 1940, and Ashton and Ashton, 1988; for Kentucky see Barbour, 1971; for Louisiana see Dundee and Rossman, 1989; for Texas see Dixon, 1987, and Garrett and Barker, 1987; for Virginia and the Carolinas see Martof et al., 1980; and for West Virginia see Green and Pauley, 1987). Books that will include extensive data on the biology of aquatic amphibians are in progress for the states of Tennessee and Virginia. In addition, separate herpetological bibliographies are available for the states of Florida (Enge and Dodd, 1992), Tennessee (Redmond et al., 1990), and Virginia (Mitchell, 1981).

Relatively recent concern for individual species has resulted in a series of books and journal articles which include a status assessment of actually or potentially imperiled aquatic amphibians. Reviews are available for West Virginia (Pauley and Canterbury, 1990), Virginia (Linzey, 1979; Pague and Mitchell, 1987; Terwilliger, 1991), North Carolina (Cooper et al., 1977), South Carolina (Harrison et al., 1979), Tennessee (Echternacht, 1980), Kentucky (Branson et al., 1981), Florida (McDiarmid, 1978; Moler, 1992d), Alabama (Mount, 1986c), and Arkansas (Reagan, 1974). Ashton (1976) provided a national checklist of amphibians and reptiles in need of conservation, and Bury et al. (1980) summarized the status of amphibians throughout the United States.

In generally assessing historical information about amphibians residing in the Southeast, endangered and threatened amphibian accounts were written by individuals familiar with the biology of the species (see Table 2). However, few assessments were based on thorough studies and none included long-term quantitative data. Symposia have seemed to highlight more of what was not known about a species than what was known. Edited proceedings have usually contained information on life history, distribution, status, and threats, whereas journal articles contained little background data.

Concern about the status of particular herpetofauna species or communities also has resulted in many inventory programs, but much of this information remains unpublished and generally unavailable. For example, intensive herpetofaunal inventories based on quantitative sampling were prepared for Lake Conway, Florida (Bancroft et al., 1983), the proposed Cross Florida Barge Canal route in the Ocala National Forest, the proposed phosphate mining area in the Osceola National Forest, and the St. Marks National Wildlife Refuge in Florida. Unfortunately, reports of the results of such surveys are difficult to obtain and often lack crucial details concerning site locations, sampling methods and intensity, and statistical analysis. Herpetofauna inventories of various other national forests (e.g., Pearson et al., 1987), military reservations (e.g., Williamson and Moulis, 1979), and state and private lands also exist for areas scattered throughout the Southeast. Herpetofaunal inventories are presently under way at Eglin Air Force Base, Florida; Ft. Stewart, Georgia; and Camp Blanding, Florida.

Recent examples of single species amphibian surveys include Rana capito in North Carolina (Braswell, 1993); R. capito and Notophthalmus perstriatus in Florida (Franz and Smith, 1993); N. perstriatus in Georgia (Dodd and LaClaire, 1995); and Ambystoma cingulatum in Florida (Palis, 1992, 1993) (also see Table 2). All southeastern states now have Natural Heritage Programs to assemble data on declining species. Some of these programs are well advanced in data analysis (e.g., Florida), whereas others are just getting started (e.g., Georgia).

In the 1970s, ecological studies generally became much more intensive and quantified, and often integrated field and laboratory work to examine hypotheses of species interactions. Although they did not initially begin as monitoring studies, the ecological studies at the Savannah River Site (SRS) in South Carolina (see Gibbons and Semlitsch, 1991, and references therein) and Hairston’s studies of terrestrial salamander competition (Hairston and Wiley, 1993) in the southern Appalachians are the only studies with truly long-term continuous data sets in the Southeast. Only the SRS study has data on all of the aquatic amphibian species in the local community.

Other studies are available covering a shorter time span. Dodd (1992) systematically monitored the amphibian community at a temporary pond in north Florida sandhills from 1985 to 1990. H. Mushinsky (University of South Florida) and A. F. Scott (Austin Peay State University) have quantitatively monitored the amphibian community on a Florida sandhill and in north-central Tennessee, respectively, since the early 1980s, although these data are not yet published. Delis (1993) compared amphibian community changes from the 1970s to the 1990s in an area of west-central Florida undergoing urbanization. Other studies have monitored a single species in a region or at a single locality for various amounts of time (M. Bailey, Alabama Natural Heritage Program; J. Palis, Jonesboro, Illinois; W. Seyle, U.S. Army Corps of Engineers, all pers. comm.), but the results of such monitoring are generally not available.

In conclusion, the literature on potentially imperiled amphibians in the Southeast is scattered and based on few quantitative studies. Much information remains unpublished or is otherwise unavailable. Therefore, it is often impossible to assess the accuracy or thoroughness of completed work. The only long-term data set on continuously monitored aquatic-dependent amphibians in the Southeast is available from SRS. At SRS, much annual variation occurs in the number of reproductive adults visiting a wetland. Reproductive output also varies annually, even when substantial numbers of adults reproduce (Pechmann et al., 1991).

There are numerous published studies on the ecology of individual aquatic-dependent southeastern species or groups of species. Locations used for such studies could serve as monitoring sites to assess the status of the species and habitat since the original studies were completed. Examples of some original assessments include the work of Trauth et al. (1992) assessing the status of an Arkansas population of hellbenders (Cryptobranchus alleganiensis) previously studied in the mid-1980s and Dodd’s (1991) study of the Red Hills salamander (Phaeognathus hubrichti). Few follow-up assessment studies have been undertaken.

Threats To Southeastern Aquatic Species

Amphibians that depend on aquatic environments in the Southeast potentially are vulnerable to a great variety of threats, although few detailed studies have specifically considered such problems within the region. The integrity of both aquatic and terrestrial habitats is important to amphibian survival, even among species that never venture beyond a single habitat type. Furthermore, the various life history stages (eggs, larvae, young, adults) may be differentially susceptible or sensitive to environmental perturbations. Studies that assess only one phase of a species’ life cycle (e.g., surveys only of breeding habitat) may overlook important ecological requirements of other life history phases. Although we tend to discuss conservation in terms of individual species, an ecosystem approach that is sensitive to all life history phases is necessary to ensure the habitat integrity that ultimately will continue to support individual species.

Literature references to southeastern aquatic-dependent amphibians that currently might be in need of some degree of management are provided in Table 2. Habitat destruction and alteration are the most commonly identified factors affecting species’ status. There are many cases where a species appears rare but is geographically peripheral to the region in question, or its true status is unknown. There is only one case where a "mysterious" decline may have occurred regarding an aquatic species. The salamander Desmognathus auriculatus appears to have declined or disappeared from sections of the Atlantic Coastal Plain in South Carolina and peninsular Florida (S. Christman, Quincy, Florida; J. Harrison, Charleston Museum, both pers. comm.), but no systematic surveys have been undertaken. However, populations of coastal plain desmognathine salamanders are known to fluctuate substantially in numbers from one year to the next (B. Means, Coastal Plains Institute, pers. comm.). Some specific threats to aquatic amphibians are discussed briefly below.

Habitat Destruction and Alteration

Even before the arrival of Europeans, Native Americans exerted considerable influence upon southeastern landscapes. Villages formed in circular patterns were interconnected by corridors and surrounded by considerable amounts of buffer land used for hunting (Hammett, 1992). Lands were used for agriculture and large areas were burned to clear land and to improve hunting. After colonization by Europeans, land clearing and ecosystem modification accelerated and have culminated in the present frenzy to redesign the landscape.

The Southeast has been rapidly increasing in human population for several decades, and its metropolitan areas are among the fastest growing population centers in the United States. In Florida alone, where more than 9 million acres (3,642,300 ha) of wetlands already have disappeared (Cerulean, 1991), the population increases by a net 900 newcomers each day. In Arkansas, 6 million of the original 10 million acres (2,428,200 of 4,047,000 ha) of Mississippi Delta wetlands have been converted to agricultural land (Smith et al., 1984). In a west-central peninsular Florida study, species richness was less in urbanized areas than in nearby pristine areas, and temporary pond breeding species disappeared entirely from the urbanized site (Delis, 1993). Although vast areas have been cleared in the Southeast for agriculture, industry, and urban use, there is virtually no assessment of the landscape effects of land conversion on amphibian populations. It seems evident, however, that habitat changes (see papers in Hackney et al., 1992; Boyce and Martin, 1993), and with them changes in aquatic amphibian populations, have been enormous.

Habitat alteration may occur without obvious large-scale topographic changes. For example, a massive boom in human population on the Edwards Plateau of Texas has increased the withdrawal of ground water from the Edwards Aquifer. As more and more water is withdrawn, water tables have decreased. In the future, springs and streams in this region are likely to dry completely, especially during periods of drought. This situation could lead to the loss of a unique aquatic biota that includes spring and cavernicolous salamanders (U.S. Fish and Wildlife Service, 1984a; Chippindale et al., 1993).

Habitat Fragmentation

Although habitat fragmentation affects biota in different ways (e.g., Mader, 1984), land use patterns resulting in fragmentation can influence amphibian population genetic structure (Reh and Seitz, 1990). Amphibian populations are most abundant when there is a mosaic of habitats located within a regional landscape (Mann et al., 1991). In such a context, metapopulations may develop which result in a dynamic equilibrium through time. However, if populations become overly fragmented, emigration and immigration may be inhibited or stopped, thus preventing recolonization from source populations. The effect of fragmentation on amphibians depends on the degree of isolation (Sjogren, 1991a). Small, isolated populations are particularly susceptible to environmental perturbations (Sjogren, 1991b) and to stochastic variation in demography that can lead to extinction even without external perturbations (Lande, 1988; Pimm et al., 1988). Isolation by habitat fragmentation thus becomes a threat to the regional persistence of species.

Forestry Practices

Most discussions of the effects of forestry on amphibians in the Southeast focus on salamanders in clearcuts (Blymer and McGinnes, 1977; Ash, 1988; Dodd, 1991; Petranka et al., 1993; Ash and Bruce, 1994; Petranka, 1994), although a few recent studies have examined amphibian communities in the coastal plain (Phelps, 1993; Dodd, 1995b; O’Neill, 1995; Phelps and Lancia, 1995; Means et al., 1996). Clearcutting reduces salamander populations because it eliminates shade, reduces forest litter (especially if litter is piled and burned), increases soil temperature, reduces soil moisture, and destroys wetlands. Herbicides are frequently used in such operations, yet little is known of their effects on amphibians (but see Bidwell and Gorrie, 1995).

Depending on the type of site preparation, clearcutting practices also reduce or eliminate burrows and other hiding places needed by aquatic habitat-dependent amphibians when they are away from their breeding sites. For example, clearcutting an area adjacent to an Ambystoma talpoideum breeding pond in Louisiana lowered the survivorship of immigrating adults using the clearcut site and displaced other adults to less suitable habitat (Raymond and Hardy, 1991). Other attributes which affect amphibian persistence after timber cutting include the status of amphibian populations prior to cutting, the type of cut (selective vs clear), the type of forest replanted, the size of the cut, the amount of time since last cut (Grant et al., 1994), and the distance to the nearest source populations. Mature southeastern pine plantations also support far fewer amphibians than adjacent deciduous forests (Bennett et al., 1980). In Florida sand pine scrub, clearcutting seems to mimic intensive wildfire; the richness and diversity of amphibians appear more dependent on the nearest water source for breeding than on the type of disturbance (Greenberg, 1993). However, clearcutting reduced amphibian species abundance in pine flatwoods tenfold by adversely affecting reproductive success (Enge and Marion, 1986). Regarding the effects of timbering, stream-dwelling species and their larvae have received little attention in the Southeast, although adverse effects to stream-dwelling amphibians caused by logging in the Pacific Northwest are well documented (Bury and Corn, 1988; Corn and Bury, 1989).

On the southeastern Coastal Plain, vast pine plantations have replaced the native longleaf pine (Pinus palustris) savanna. During planting and site preparation, much of the land was ditched in an effort to speed water runoff. Literally thousands of acres of wetlands disappeared or were substantially altered. Ditching occurred between ponds to facilitate water transfer; water essentially flowed downhill, although slowly, thus reducing available hydroperiods for amphibian larval development. A second type of ditching occurs around wetlands. Circumferential ditching results in lowered water tables with concomitant vegetative changes, thus drastically altering or eliminating hydroperiods. Unditched ponds are more persistent than ditched ponds, and have greater amphibian species richness during dry periods (Harris and Vickers, 1984; Vickers et al., 1985). In addition, more aquatic amphibian species are associated with unditched ponds. This is especially important because many temporary pond-breeding amphibians exhibit breeding site fidelity and other obligate breeding requirements which can be impacted by ditching.

The loss of the longleaf pine forest on the coastal plain of the southeastern United States has been dramatic (Means and Grow, 1985; Noss, 1989; Boyce and Martin, 1993; Stout and Marion, 1993; Ware et al., 1993). Concern for the survival of the coastal plain forest in Georgia was expressed at least as early as 1906 because of logging, turpentining, and land clearing for agriculture and "civilization" (Harper, 1906). Since the 1940s, old-growth longleaf pine forest has been converted to slash (P. elliottii) and loblolly (P. taeda) pine plantations throughout the Southeast. In southeast Georgia, for example, longleaf pine declined 36 percent between 1981 and 1988 to 230,000 acres (93,081 ha; see Johnson, 1988), whereas in southwest Georgia, longleaf pine declined four percent during these same years to 205,000 acres (82,963 ha)(Thompson, 1988). Today, less than one percent of the old growth longleaf pine forest remains (of the more than 70 million acres [28,329,000 ha] present when Europeans colonized the continent). Most remaining forest is scattered and poorly managed. Even in national forests, longleaf pine has declined substantially (Means and Grow, 1985). In the last few decades, drastic changes probably have occurred in the composition and structure of the amphibian community in regions that formerly held longleaf pine (Dodd, 1995b; Means et al., 1996), but no baseline data exist to document the effects of this continuing massive landscape alteration.

Mining

Extensive coal strip mining is carried out in West Virginia, Virginia, Kentucky, Tennessee, and Alabama. In many instances, mining occurs directly through small streams or ponds, and mine tailings are pushed into the larger rivers. In Florida, vast areas have been strip-mined for phosphate. Mining not only destroys aquatic amphibian habitats outright, it also results in toxic pollution, decreased pH, and siltation of streams and rivers. Low pH combined with high levels of conductivity (an indication of the presence of pollutants) limit the presence of larval salamanders of the genus Desmognathus in mine-affected streams of the Cumberland Plateau (Gore, 1983). Paradoxically, amphibians have bred in strip mine ponds as long as the pH was not too low and toxic waste was prevented from entering the pond (Turner and Fowler, 1981; Lacki et al., 1992).

Transportation Corridors

Transportation corridors, especially roads, can have serious deleterious effects on amphibian populations (Langton, 1989). Road construction can lead to habitat destruction in both terrestrial and aquatic environments, and can negatively alter breeding habitats through increased siltation. Increased siltation can lead to increased amphibian mortality because of its own secondary effects. For example, nearly all aquatic life was eliminated downstream after U.S. Highway 441 was rebuilt in 1963 in the Great Smoky Mountains National Park. Toxic substances associated with leachates from roadfill were suspected as the cause. Laboratory experiments confirmed that roadfill leachates were toxic to larval shovel-nosed salamanders (Leurognathus marmoratus). The major components of the leachate responsible for toxicity included low pH combined with high heavy metal concentrations (Mathews and Morgan, 1982).

Roads may separate overwintering sites from breeding sites and increase mortality as animals attempt to cross. For example, Heine (1987) demonstrated that 26 vehicles per hour on one road was enough traffic to ensure that no toads successfully crossed. Road construction also can lead to habitat fragmentation, and in doing so can hinder immigration and emigration, and isolate populations (Laan and Verboom, 1990) leading to deleterious effects associated with small population size (Sjogren, 1991b). Furthermore, the noise levels and artificial lights associated with traffic may disrupt breeding activities. Noise makes it difficult to hear conspecifics or causes frogs to completely stop calling (author’s pers. obs.). Bright artificial lighting can adversely affect frogs’ abilities to detect and consume prey (Buchanan, 1993).

Climate Change

If climate changes, possibly in response to increasing levels of greenhouse gases, then there are bound to be changes in the diversity of southeastern amphibians. Most of our endemic species and species-rich amphibian communities are found on the higher elevations of mountains in the cool southern Appalachians and Ozarks, or in specialized coastal plain habitats, such as temporary ponds. Spring adapted salamanders of the Edwards Plateau are sensitive to alterations in ground water levels. These species would be particularly susceptible to climate changes which alter rainfall patterns or elevate mean annual temperatures. However, the potential for changes in amphibian diversity seems to have been overlooked in the climate change debate. For example, amphibians are mentioned, briefly, only twice in 26 chapters of a recent book which examines the effects of global warming on biological diversity (Peters and Lovejoy, 1992).

pH

The acidity of aquatic habitats can play a major role in limiting the distribution of amphibians. Decreased levels of pH in aquatic habitats may result from acidic precipitation or point-sources of pollution, such as abandoned mines. Acid concentration may increase steadily or come in pulses, such as during heavy rains or from snow melt. Although the Southeast has not experienced as many problems from acid rain as other parts of the United States, the acid content of our precipitation is increasing (Haines, 1979). For example, H+ increased 19-fold from 1955 to 1979 in Great Smoky Mountains National Park (Mathews and Larson, 1980). Bioassay results suggested that pH levels were near toxic to larval shovel-nosed salamanders, although not as toxic to adults (Mathews and Larson, 1980).

The literature on the effects of pH on amphibians is voluminous and complex (Freda, 1986; Dunson and Wyman, 1992). Low pH has different effects on different species of amphibians and, indeed, there may be intraspecific differences in pH sensitivity that varies geographically. Furthermore, these intraspecific differences may or may not have a genetic basis (Pierce and Wooten, 1992). In general, the eggs and developing larvae are the most sensitive life stages to low pH (< 4.5). A low pH alters the cellular chemical environment by disrupting the Na+ and Cl- balance both in terrestrial (Frisbie and Wyman, 1991) and aquatic life stages (Freda and Dunson, 1984). This, in turn, affects salamander spatial distribution since salamanders avoid soils of low pH (Wyman and Hawksley-Lescault, 1987; Freda and Taylor, 1992). Low pH also may impair the vitally important chemosensory system of amphibians (Griffiths, 1993) and inhibit larval feeding (Roudebush, 1988).

Low pH can also have indirect effects which can kill eggs, larvae, or even adults (Sadinski and Dunson, 1992). A low pH acts to inhibit amphibian egg capsule enlargement, and thus limits the space available to the growing embryo. In addition, high acidity inhibits proper jelly formation. Jelly allows spacing of the eggs within an egg mass which ensures that each developing embryo has an adequate oxygen supply (Seymour, 1994). If jelly does not form properly, death from anoxia results. Chronic or intermittent low pH also can disrupt environmental trophic interactions (Sadinski and Dunson, 1992), and can lead to problems associated with long-term environmental stress. For example, phytoplankton which are fed upon by tadpoles are also sensitive to low pH (Haines, 1981).

Toxic Substances

A great many substances are likely toxic to amphibians, at least during part of their life cycle. Toxicants need not be synthetic chemicals. For example, salt spread on roads during winter can affect the chemistry of amphibian breeding sites. Toxic chemicals can enter the environment in many ways, both intentionally and accidentally. There have been numerous instances of inadvertent release of toxic materials into aquatic habitats because of highway or railroad accidents.

Surprisingly little research has been done on the effects of toxic chemicals on amphibians, and even then most work has focused on only one part of the life cycle. Examples of toxic materials known to adversely affect amphibians include heavy metals (aluminum, mercury, selenium), pesticides (toxaphene, heptachlor, malathion, endrin, methoxychlor), herbicides (DEF, trifluralin, atrazine), fungicides (furanace, malachite green), phenols, carbon tetrachloride, and nitrite. Literature summaries are provided in Birge et al. (1980), Power et al. (1989), and Hall and Henry (1992).

Data on the level of toxic chemicals in wild populations of amphibians, much less those of the Southeast, are nearly non-existent. However, Hall et al. (1985) noted metabolites of DDT as well as PCBs (primarily chlordane constituents) in Necturus lewisi from the Tar and Neuse rivers, North Carolina. The herbicide atrazine was implicated as contributing to large frog (Rana pipiens) die-offs in Wisconsin (Hine et al., 1981).

In addition to direct effects, certain toxicants may affect amphibians differently depending on pH. For example, aluminum has adverse effects upon amphibians, but the level of adversity differs depending on species, life stage, and pH (Beattie and Tyler-Jones, 1992; Bradford et al., 1992; Jung and Jagoe, 1993). Lowered pHs amplify the toxicity of heavy metals to amphibians.

Endocrine Mimics

Many chlorinated chemicals (DDT, PCBs, etc.) have been dumped in huge quantities into the environment during the 20th century, and as they travel throughout food chains they become magnified in concentration. Chlorinated chemicals can act to impair development, block intracellular communication, and induce enzymes that break down hormones. In addition, many of these persistent compounds function, even in minute quantities, as hormones, especially mimicking estrogen. Some of the side effects of endocrine mimics are thyroid dysfunction, metabolic abnormalities, decreased fertility, birth deformities, abnormal sexual development, and immunosuppression (Carey and Bryant, 1995; Stebbins and Cohen, 1995). Although no specific examples exist yet for amphibians, xenobiotics have been implicated in partial sex reversals and gonadal feminization in a wild Florida population of American alligators (Alligator mississippiensis) (Guillette et al., 1994).

Amphibians are likely to be especially sensitive to the action of endocrine mimics because they are in close direct contact with chemicals in their environment, and the amphibian skin and egg capsule are highly permeable. Because hormones normally function in minute quantities and are vital to normal development (Hourdry, 1993), susceptibility to xenobiotics could be devastating during the complex changes that occur during hormonally-induced amphibian metamorphosis.

Ultraviolet-B Radiation

Recent evidence suggests ultraviolet-B (UV-B) radiation has adverse effects on amphibian larval hatching success and that sensitivity to UV-B varies among species (Blaustein et al., 1994a) or is exacerbated by low pH (Long et al., 1995). Species with high levels of photolyase (e.g., Pseudacris spp.), an enzyme involved in DNA repair of ultraviolet radiation damage, are less prone to the adverse effects of UV-B radiation than species with low levels of photolyase (e.g., Bufo spp., Rana spp.). Many populations of Bufo spp. and Rana spp. have declined in the western United States, whereas Pseudacris triseriata populations have not. Frog embryos (Rana clamitans and R. sylvatica) exposed to high levels of UV-B had higher rates of developmental abnormalities and increased mortality than controls which were shielded from UV-B (Grant and Licht, 1993). UV-B also can have detrimental effects on embryo growth. UV-B radiation has increased recently in the northern hemisphere because of ozone depletion (Blumthaler and Ambach, 1990; Kerr and McElroy, 1993). If UV-B adversely affects southern Appalachian anurans, toads (Bufo spp.) and true frogs (Rana spp.) would seem most likely to be affected.

Exotics, Predators, and Competitors

There is no literature on the effects of the many exotic fishes in southeastern waters on native herpetofauna. Fish may be both competitors and predators of amphibians, depending on life cycle stage (Bristow, 1991). They have been implicated in declines of western amphibians both as predators (Bradford, 1989) and as disease vectors (Blaustein et al., 1994b). Stocking of predatory fishes in ponds previously free of fish undoubtedly leads to a change in the amphibian community because many amphibians are defenseless against fish predators. Conversion of temporary ponds to permanent ponds by digging and blasting, followed by fish introductions, often leads to a loss of the temporary pond breeding species. The effects of exotic frogs, especially the marine toad (Bufo marinus) and Cuban treefrog (Osteopilus septentrionalis), on native amphibians are unknown, although anuran species richness was reduced in at least one area having marine toads, compared to a similar area without them (Rossi, 1981). Release of other exotics undoubtedly occurs with unknown effects. One south Florida tropical fish dealer reported selling 50,000 eastern newts in Florida that originated from outside the state (Enge, 1991). Many of these exotic newts undoubtedly were released intentionally or unintentionally.

Birds and mammals also may exact a substantial toll on amphibian populations, especially exotic cattle egrets, armadillos, and wild hogs. In addition, populations of some native species, such as raccoons, may become so large because of a lack of natural predators and adaptation to human surroundings that they in turn reduce amphibian populations beyond normal levels. The overabundance of some native species is an issue which biologists are only beginning to confront (Garrott et al., 1993).

Finally, there are few data on the effects of exotic invertebrates, especially imported red fire ants (Solenopsis invicta), on native amphibians. Ground-dwelling vertebrates are especially sensitive to this ravenous predator (Mount, 1981), and fire ants have been reported to kill endangered Houston toads (Bufo houstonensis) as they metamorphose (Freed and Neitman, 1988). Fire ants are especially abundant in the moist perimeter surrounding ponds and lakes, and they can float in mats across ponds from vegetation clump to vegetation clump. Fire ants have few predators and have expanded their range throughout the Southeast.

Collecting

Collecting specimens for the pet trade or biological laboratories probably has had some impact on local amphibian populations, but few data are available. Trauth et al. (1992) suspected that collection of hellbenders in the Spring River, Arkansas, contributed to observed population declines. From 1 July 1990 to 30 June 1991, 804 salamanders and 18,170 frogs were collected legally for the Florida pet trade (Enge, 1991). Included were 5,066 Hyla cinerea, 3,265 Bufo terrestris, 2,674 Hyla gratiosa, and 249 Siren lacertina. In 1992, 246 salamanders and 23,019 frogs were collected and sold in the pet trade in Florida (Enge, 1993). Most sales went to New York, Pennsylvania, and Tennessee. Concern for the effect of biological supply house collection on frog populations is not new (Gibbs et al., 1971). In the early 1970s, U.S. frog suppliers shipped 9 million frogs (over 326,000 kg) per year. The number of frogs shipped by southeastern supply houses is unknown.

Loss or Decline of Associates

If species that are preyed upon by amphibians decline or disappear, amphibian populations may be expected to follow suit. The use of pesticides and the influence of toxics, pH, and habitat alteration all may be expected to affect amphibian prey populations. In addition, amphibians sometimes rely upon the burrows of other species for shelter when they are away from ponds. If these associated animals are eliminated, fewer shelters may be available. A few amphibians inhabit the burrows of specific associates. For example, gopher frogs (Rana capito) nearly always reside in sympatric gopher tortoise (Gopherus polyphemus) burrows when the frogs are not at breeding ponds. Yet, the number of gopher tortoises is estimated to have declined by 80 percent during the last 100 years (Auffenberg and Franz, 1982). The effect of the decline of tortoises and their sheltering burrows on gopher frogs is unknown.

Drought, Cold, and Disease

Drought, cold, and disease are natural factors that affect amphibian communities (e.g., Dodd, 1993, 1995a). Drought can lead to localized extirpation. Excessive cold can induce winterkill in torpid amphibians. Disease can wipe out populations. However, the chronic effects of these factors on amphibian populations, if any, remain unknown. Under pristine conditions, amphibian populations often may expand and contract in response to such natural variables affecting local distribution, thus forming a dynamic equilibrium (Sjogren, 1993a, 1993b). Under present human-dominated landscapes, however, populations may be so fragmented or under such a variety of stresses that they are unable to rebound from extrinsic environmental factors causing periodic population fluctuations. If many amphibian populations function as metapopulations, the long-term survival of local populations might be jeopardized by isolation from source populations coupled with natural environmental fluctuations.

In fact, "natural" factors may not be as natural as they first appear. For example, droughts may result from global climate change or they may be magnified by habitat alteration such as deforestation or overgrazing. The effects of disease also might be facilitated by human activity. Carey (1993a, 1993b) has proposed a model whereby sublethal stress (such as that associated with chronic low but sublethal pH, or high concentration of a toxicant, or increased UV-B radiation) induces either direct or indirect immunosuppression because of the prolonged elevation of adrenal cortical hormones. Depressed immunity makes the animal more prone to naturally occurring pathogens, such as red-leg disease causing bacteria (Aeromonas spp.), especially during periods of torpor. This model is consistent with observations on declining amphibians in many Rocky Mountain populations where amphibian populations have been known to decline gradually and then one year simply fail to emerge from hibernation.

A pathogenic fungus has been implicated recently in the decline and disappearance of Bufo boreas in the western United States (Blaustein et al., 1994b). The fungus (Saprolegnia ferax) is circumglobal in distribution and commonly occurs on fish. However, fish are not native to the high mountain habitats occupied by B. boreas, and the pathogen is thought to have been introduced when trout and salmon were stocked in high mountain streams and lakes. The same fungus has extirpated other frog populations in the U.S. and Europe (for a review of this topic see Blaustein et al., 1994b). Although the extent of amphibian fungal infections is unknown in the Southeast, every egg mass (Rana sp.) I examined during March 1994 in several ponds on Trail Ridge in southern Georgia was infected by an as yet unidentified fungus.

Summary

The southeastern United States holds a rich temperate amphibian assemblage containing a great degree of endemism. Endemic species are especially well represented among the salamanders. A varied topography and complex geologic history have provided the necessary conditions that have resulted in this region’s high level of speciation. However, the amphibians of this area, and particularly the fully aquatic species, face a multitude of threats to their long-term existence. These threats generally do not act independently, but instead act in concert to have potentially serious long-term effects. Many amphibian species have been identified as needing conservation programs and management, but few scientific studies have assessed direct effects to species or ecosystems. There also are few studies detailed enough to show trends or to separate unnatural trends from normal population fluctuations.

Although natural population fluctuations undoubtedly exist, it is extremely naive and certainly not objective to call simply for "more monitoring" (Pechmann and Wilbur, 1994). At a time when conservation related funding is nearly nonexistent and no agencies seem able to initiate long-term monitoring on a scale required to assess wide-ranging threats to amphibians, the call for more monitoring seems an effective mask for doing nothing. How can interest be generated in monitoring "common" or non-threatened species, much less communities and ecosystems, when programs directed at the conservation of critically endangered species are under-funded or not funded at all? Given the cumulative assaults on the biosphere in the late 20th century, I suggest Chicken Little is better in tune with biological and political reality than Nero with his fiddle (see Blaustein, 1994). Rome, after all, burned.

As Gibbons (1988) has discussed, a new attitude is needed toward the recognition of the importance of amphibians to ecosystem functioning. No longer can these species be assigned a role of non-importance in wildlife and land management. Attention must be focused on threats to species inasmuch as these threats may be symptomatic of serious environmental problems. We need to study the seemingly common species (Dodd and Franz, 1993), as well as the rare or endangered species. Our casual perceptions may not always give an accurate assessment of population status. Finally, we need an ecosystem, landscape, and watershed approach to understanding the role of amphibians in imperiled aquatic systems as well as adequate funding from private and governmental agencies (Mittermeier et al., 1992) to carry out necessary research and management programs.

Acknowledgements

I thank Ronn Altig, Dick Franz, Marian Griffey, and Stan Trauth for comments on the manuscript. Members of the southeastern Declining Amphibian Populations Task Force supplied some information used in this paper. I thank Dave Collins for inviting me to participate in the important imperiled aquatic fauna symposium. This paper is dedicated to the late Roger W. Barbour, who introduced me to the wonderful world of salamanders during my undergraduate days at the University of Kentucky.

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