Plight of the Salamanders

Introduction to the Issue:

                Salamander species across the globe are facing serious threats to their survival. This plight is affecting many different species and has many facets and causes adding to its complexity. Salamanders face serious declines from a deadly epidemic stemming from a fungus known as Batrachochytrium dendrobatidis. Also, the current global climate and environmental changes, especially ones associated with humans, have led to reductions in suitable habitat for salamander populations. Human population growth and land use has greatly expanded over time have led to habitat fragmentation in addition to reduction. In addition to direct habitat loss from humans, salamanders are suffering from losses to better suited species that are shifting their ranges. Invasive species also threaten the genetic purity of native salamander species when they through interbreeding, changing the genetic diversity of the population. Once again, this problem can be traced back to humans who are most often the means by which the invasive species enter the new system. The global environment is changeable and is shifting both in terms of climate an in the way that its inhabitants interact with one another. Humans, especially, are effecting salamander populations and should be conscious of their effects.

                Part of why humans are less aware than they should be about this problem, is that amphibians and especially salamanders are mysterious creatures. Often the only times they can be studied or even seen by humans is when they are migrating during the breeding season. This means that many humans don’t get the chance to interact with salamanders and thus cannot be expected to care about the survival of salamander populations. Conservationists have attempted to remedy these inconsistencies by forming the Declining Amphibian Populations Task Force in 1991 (“Declining Amphibian…). This task force was established by the Species Survival Commission and the World Conservation Union (IUCN) to more consistently collect information about amphibians and the possible causes of their declines (“Declining Amphibian…). This parallels other conservation efforts, many of which have their main goal as gaining a more comprehensive knowledge of the species that are in danger. One way they are accumulating knowledge is through in depth research and field tracking to understand and fill in the gaps that are rampant in many salamander species’ life histories. Thankfully, it seems like humans finally comprehend the urgency to this plight and have begun to respond.

Problems Associated with Limited Range:

                Many salamander species are suffering from range reductions. But what happens to salamander species that start off inhabiting a very small range? Certain habitats can lead to species maintaining very small ranges. One common area that promotes small ranges is montane regions that include many variations associated with elevation changes, such as vegetation type and temperature (Bayer et al. 2012). Often what happens in these regions is that mountain top habitats encourage allopatric speciation events (Bayer et al. 2012). One example of this a recently discovered species, Plethodon sherando, which evolved in a very small range in the higher areas of the Blue Ridge Mountains of Virginia (Bayer et al. 2012). This species is surrounded by a closely related species, Plethodon cinereus, which comes into contact to P. sherando and may lead to conservation issues (Bayer et al. 2012). A general problem associated with living in high elevation habitats is that species cannot simply shift their range northwards or upslope to help respond to climate changes because eventually the mountain will end (Bayer et al. 2012). Studies have modeled what could happen under climate change projections to these types of species and they generally predict range contractions or even extinctions to these mountain top species (Bayer et al. 2012). Mountain top species have to worry about not being able to shift ranges, and about being replaced with competing species that possess higher thermal tolerances (Bayer et al. 2012). These competing species are ones that often live lower in elevation and need to shift their ranges upward because of encroachment from other species. Another problem can occur if the species hybridize with surrounding species (Bayer et al. 2012). This may lead to genetic swamping, in which the original species’ gene pool is overrun by the surrounding species’ genes and they cease being distinct populations (Bayer et al. 2012).

                In this specific case, P. sherando occupies a range of less than 80 km² along the top of the Big Levels area of Augusta County in Virginia (Bayer et al. 2012). It has been found on slopes down to 640-670 m where there is a drastic transition to P. cinereus leading to the main contact zone (Bayer et al. 2012). However, there is an additional contact zone at higher elevations where Big Levels wraps around to reach the main Blue Ridge (Bayer et al. 2012). There is not much known about P. sherando which means it has not been classified for conservation efforts, and thus not considered in planning for the region (Bayer et al. 2012). In the spirit of gaining more knowledge about lesser known species, Bayer et al. (2012) set out to determine the results of the contact zone between P. sherando and P. cinereus. Figure 1 displays the study zone and includes the contact zones between the two species as well as where P. sherando occurs alone.

Figure 1: Filled circles represent P. sherando only, while crosses indicate sties with P. cinereus only. CZ1 and CZ2 are contact sites where both species are found (Bayer et al. 2012).

             Their study proved the genetic uniqueness of P. sherando in addition to the already known morphological differences between P. sherando and P. cinereus (Bayer et al. 2012). Figure 2 is the compilation of their genetic findings and demonstrates that under all of their testing, P. sherando and P. cinereus remain monophyletic and thus distinct species (Bayer et al. 2012). 

Figure 2: This phylogeny demonstrates the genetic uniqueness of P. sherando and P. cinereus (Bayer et al. 2012).

Because P. sherando is genetically unique from P. cinereus, it is likely that it evolved during a period of isolation on the mountain top (Bayer et al. 2012). However, despite its distinct genetics, P. sherando is still threatened because of its limited range and because its range is surrounded by P. cinereus. Plethodon salamanders are known to frequently hybridize and, despite being unique species, P. sherando and P. cinereus may still hybridize leading to issues with genetic swamping of the P. sherando gene pool (Bayer et al. 2012). Bayer et al. (2012) did not determine this problem to pose immediate danger to P. sherando, but this study did have limitations so more research may be warranted before dismissing this issue (Bayer et al. 2012). Climate modeling suggests that a bigger issue will be climate change, leading to competition over range (Bayer et al. 2012). Because P. cinereus is found at warmer, lower elevations, it is likely that it may expand its range and outcompete P. sherando (Bayer et al. 2012). But this is not certain because P. cinereus has a larger range which includes high and low elevation areas so warmer conditions from global warming may not favor P. cinereus (Bayer et al. 2012). Because of these findings, in addition to the fact that the other four mountaintop salamanders in the region have special status listings, Bayer et al. (2012) suggests that P. sherando be listed and obtain a formal conservation status. This study represents the benefit that additional research can have on lesser known salamander species and how once more is known, humans can then use this information to make decisions that will help, rather than harm, salamander species.

                 A different study dealt with the seepage salamander, Desmognathus aeneus, which also live in relatively small ranges, and found good news in terms of this species classification. Graham et al. (2012) looked at D. aeneus to determine how current populations compared to historic ones, and whether any differences would warrant new federal statuses. D. aeneus is found in disjunct populations throughout the Blue Ridge, Coastal Plain, and Piedmont provinces of the Southeastern United States (Graham et al. 2012). D. aeneus was recently petitioned for federal protection under the Endangered Species Act because of its limited range, presumed strict habitat requirements, and habitat modification all of which threaten this species’ survival (Graham et al. 2012). Graham et al. (2012) completed 136 surveys at 101 sites, 46 of which had historical data to compare to current findings. These surveys included one to three people combing through the likely areas which included hardwood forests and stream or seepage margins (Graham et al. 2012). However, there is some habitat distinction between the northeastern populations and those in the southwest (Graham et al. 2012). Populations in the northeast are more terrestrial and live between the leaf/leaf mold layer, and the underlying soil, as well as under rocks, coarse woody debris, or moss mats on boulders in shady hardwood or mixed forests (Graham et al. 2012). In contrast, populations in the southwest are also abundant in hardwood forests but associated more often with permanent seepage sites which are comparable to the moisture levels of higher elevation sites (Graham et al. 2012). The surveyors turned over coarse woody debris, raked leaf litter, and looked through moss clumps for an hour at a time measuring encounter rates at each site (Graham et al. 2012). Then they were able to compare the encounter rates to the historical data from Alabama (Graham et al. 2012).

                D. aeneus was found at 36 of the 46 historical locations (78%) and 35 of the 55 non-historical localities which represented previously unknown populations (Graham et al. 2012). The majority of the salamanders were found under leaf litter (46%) or moss mats (45%), with only 9% found under coarse woody debris (Graham et al. 2012). Figure 3 represents the typical microhabitats for D. aeneus; the mossy mats were used by larvae, juveniles, adults, and nesting D. aeneus (Graham et al. 2012).

Figure 3: These photos demonstrate some examples of microhabitats that D. aeneus prefers. Photo A represents mossy mats, while photo B represents leaf litter piles. (Graham et al. 2012).

These new sites signify range expansions and filled distribution gaps, all of which is good news for this species because the findings do not indicate range-wide population decreases (Graham et al. 2012). In fact, this evidence conflicts with similar inventories of other eastern US species that are currently declining including the southern dusky salamander, Desmognathus auriculatus, green salamanders, Aneides aeneus, and hellbenders, Cryptobranchus alleganiensis (Graham et al. 2012). Thus, it appears that most populations of D. aeneus are relatively numerous and secure, meaning that most do not require listings that would divert funds or resources from other species that are more threatened (Graham et al. 2012). Graham et al. (2012) deemed that populations in North Carolina would classify as secure if not for its limited distribution which requires that it be classified instead as vulnerable. Meanwhile in Georgia, the discovery of new localities makes it likely that more than 100 populations are maintained which leads to the recommendation that these populations have their status lowered to secure (Graham et al. 2012). Alabama populations were fewer than twenty historically, and remain around this number which classifies them as imperiled (Graham et al. 2012). The final areas studied were in Tennessee and South Carolina, which both had many undocumented populations despite Graham et al. (2012) only sampling a couple sites in each state; thus they recommended that the populations remain secure or unchanged (Graham et al. 2012). Overall D. aeneus appears to be surviving well in most of their populations, well enough to warrant a secure classification, despite similar species suffering declines. 

Human Effects within Suitable Habitats, Specifically Logging:

                Not only are salamanders often suffering from limited distributions, but they are also facing increased threats because of habitat degradation and conversion (Peterman et al. 2011). Many salamander species face huge habitat losses because of logging, and they must respond to these changes in their environment. Peterman et al. (2011) proposes that organisms respond to habitat changes in three different ways, which has led to their three hypotheses. The first hypothesis, the mortality hypothesis, implies that the organism cannot survive in the new environment and cannot successfully migrate to a more suitable habitat (Peterman et al. 2011). The results of this theory would mean reduced abundances due to deaths due to desiccation, starvation, or loss of refuge (Peterman et al. 2011). The second hypothesis, the retreat hypothesis, assumes that salamanders would retreat into underground refuges within the altered environments and display decreased activity levels (Peterman et al. 2011). The third hypothesis, the evacuation hypothesis, presumes that the salamanders would leave the unsuitable habitat for more preferred microhabitats (Peterman et al. 2011). The three hypotheses are not mutually exclusive and likely all occur in different situations of degradation as well as when dealing with different species (Peterman et al. 2011).

                Peterman et al. (2011) studied these hypotheses in headwater stream habitats which are extremely threatened features abundant in montane forests. These habitats suffered the effects of logging but there were varying amounts of riparian buffers left between the intact forest habitats and the new logged habitats (Peterman et al. 2011). Buffers ranged from 0 m, to 9 m, to 30 m, and Peterman et al. (2011) measured the amount of use and growth of salamanders in these terrestrial habitats in addition to population changes in the adjacent headwater streams. The retreat and the mortality hypotheses would be difficult to separate in the short term because both would result in decreased abundances, and terrestrial habitat use will be contracted (Peterman et al. 2011). However the evacuation hypothesis will be observed if salamanders move away from unsuitable habitats decreasing abundances, but increasing abundances in the remaining suitable portions (Peterman et al. 2011). This would mean that stream-dependent salamanders would likely move toward or into the stream and possibly moving up or down the stream to continue to find more suitable habitats (Peterman et al. 2011). The results of the three hypotheses mean different things for salamander conservation. If the salamanders respond according to the retreat or evacuation hypothesis, it may be possibly to reverse the effects of logging should the vegetation structures and microclimates be allowed to return (Peterman et al. 2011). Peterman et al. (2011)’s study is key in understanding the true effects that logging have on salamander populations.

                Their study included measuring the abundance, terrestrial habitat use, growth, in-stream capture frequency, and in-stream population abundance in comparison to the three different riparian buffers left by logging (Peterman et al. 2011). The sites were logged and then classified into the 0 m, 9 m, 30 m, and control zones; the control being areas where no logging had occurred (Peterman et al. 2011). They found that in-stream capture rates increased in all treatments over the course of a year and that the greatest increases were found in the 0 m, and 9 m treatments (Peterman et al. 2011). Figure 4 demonstrates this increase in both Desmognathus monticola and D. ocoee over the trapping period (Peterman et al. 2011).

Figure 4: Both D. monticola and D. ocoee were captured more often closer toward the streams which indicates that  salamanders may be evacuating the unsuitable habitats (Peterman et al. 2011).

Correspondingly, salamander abundances in riparian areas decreased from 71-100% throughout all the treatments (Peterman et al. 2011). However, Peterman et al. (2011) found that the different terrestrial species responded differently to the different treatments. The most terrestrial dependent species, Eurycea wilderae, had significantly contracted its abundance away from the stream (Peterman et al. 2011). While D. monticola, the least terrestrial species, had not moved significantly closer to the stream in the 9 m treatment while E. wilderae and D. ocoee, a middle species, had both moved closer (Peterman et al. 2011). Overall this meant reductions in distributions and abundances away from the streams and in the riparian habitat, as well as increases in abundances in the stream for the 0 m, and 9 m treatments (Peterman et al. 2011). Figure 5 shows the dramatic decreases in abundances in the riparian habitats (Peterman et al. 2011).

Figure 5: Huge decreases in salamanders per transect in the riparian assemblages that have been influenced by human effects (Peterman et al. 2011).

Their results indicate that salamanders are most closely following the evacuation hypothesis as salamanders are actively leaving the altered environments for the stream (Peterman et al. 2011). However it is likely that some salamanders, both at the species and individual level, are following the other hypotheses as well (Peterman et al. 2011). It is very likely that some are retreating underground or dying as a result of the riparian logging, thus supporting the other hypotheses, but to a lesser degree (Peterman et al. 2011).

                Currently, headwater habitats are not considered important in management efforts, and these results show that there are dramatic changes happening as a result of logging in the area (Peterman et al. 2011). Peterman et al. (2011)’s work shows that both the 0 m, and 9 m, buffers were not sufficient in preventing decreases in salamander abundances, and that a 30 m buffer was able to amend the effects, if not at least delay them(Peterman et al. 2011). While there may actually be a benefit to salamander species should woody debris in logged areas be allowed to decompose, this study shows that logging has hug effects of salamander species and these need to be considered when designing management plans (Peterman et al. 2011).

Invasive Species that Interbreed with Natives Creating Hybrids:

                When an invasive species invades the habitat of a native species, it can cause population declines due to limited amounts of shared resources. But this problem can be solved rather easily through removing the invasive species if declines begin to threaten the native. A more complicated problem occurs when the native and the invasive species can interbreed. There are a few different potential consequences of these hybridization events, ranging from complete fusion of the two lineages and extinction of one or both species, to only introducing a few new genes (Fitzpatrick and Shaffer 2007). One possible outcome is that the hybrids may possess lower fitness and could reduce the average population fitness, which can be a huge problem if the population is already vulnerable to extinction (Fitzpatrick and Shaffer 2007). Another is that the new hybrids may contribute to the evolution of invasiveness due to genetic swamping (Fitzpatrick and Shaffer 2007). This result would decrease the biodiversity and is often considered to make the new population less valuable because they are no longer the authentic native population because of the introduction of invasive genes into the gene pool (Fitzpatrick and Shaffer 2007).  Both of these outcomes are considered to harm the native population, but depending on the circumstances hybridization could actually be a good thing. For example if a population is so threatened that extinction is imminent, introduction of new genes may help the species by decreasing the effects of inbreeding depression or even provide key adaptations to allow the species to survive in degraded habitats (Fitzpatrick and Shaffer 2007). Thus one of the big problems that this situation brings up is under what circumstances genetic purity should be preserved and when this situation may be instrumental in the native species’ survival.

                The case study of hybridization between the native California tiger salamander, Ambystoma californiense, and the invasive barred tiger salamander, Ambystoma mavortium, demonstrates the difficulties that hybridization brings up for conservation efforts (Fitzpatrick and Shaffer 2007). These two salamanders have been geographically isolated from each other for approximately five million years but 50-60 years ago fishermen began to introduce A. mavortium larvae as bait (Fitzpatrick and Shaffer 2007). As a result, there is a budding hybrid swarm in the Salinas Valley in California (Fitzpatrick and Shaffer 2007). Currently 15-20% of A. californiense’s natural range is currently under threat from hybridization stemming from multiple introductions across the range as well as long dispersals of A. mavortium(Fitzpatrick and Shaffer 2007). This situation represents a perfect case study because A. californiense  is currently listed as Threatened under the US Endangered Species Act and can provide key evidence of the effects of hybridization (Fitzpatrick and Shaffer 2007).

              Fitzpatrick and Shaffer (2007) estimated the viability of hybrid larvae over the first few weeks of hatching because high mortality at this time provides a large selection factor for the success of the various genotypes. They looked at five different wild populations across the three main breeding habitats in the Salinas valley including natural vernal pools, seasonal cattle ponds, and perennial ponds (Fitzpatrick and Shaffer 2007). Each salamander had each nuclear marker classified as either homozygous native, heterozygous, or homozygous introduced (Fitzpatrick and Shaffer 2007). They were looking for evidence of hybrid dysfunction, in which the hybrids are less fit than the natural species, or hybrid vigor, in which hybrids are more fit than the original species (Fitzpatrick and Shaffer 2007). Results across all ponds supported hybrid vigor in that individuals with more heterozygous markers, and more mixed ancestry survived better (Figure 6) (Fitzpatrick and Shaffer 2007).

Figure 6: In all ponds gene heterozygosity and mixed ancestry were preferred and became more prevalent.

Ponds that began with a higher proportion of native genes became more mixed just as ponds with more introduced genes became more mixed which indicate that mixed ancestry is favored over either original species (Fitzpatrick and Shaffer 2007).  Overall the effects of mixed ancestry were more significant than heterozygosity which indicates that higher survival rates were not solely from heterozygote advantage but were due to hybridization (Fitzpatrick and Shaffer 2007).  However the study’s analysis revealed that both heterozygote advantage and recombinant hybrid vigor were benefitting the new hybrids which may lead to some loci being for native alleles, some for introduced alleles, and others segregating native and introduced alleles as polymorphisms (Fitzpatrick and Shaffer 2007). Hybrid vigor, in the long run, tends to maintain genetic variation, decrease linkage disequilibrium, and homogenize allele frequencies across habitats (Fitzpatrick and Shaffer 2007). There are three different process that are occurring together to bring about this stabilization. The first is genetic drift which is working to increase genetic variation in the ponds thus allowing for neither extreme to stabilize (Fitzpatrick and Shaffer 2007). Second is immigration from native ponds into hybrid zones is adding more native alleles to the zone, and effectively creating a gradient in allele frequencies (Fitzpatrick and Shaffer 2007). Finally they found that in perennial ponds, there were higher frequencies of alleles from A. mavortium which allows there to be areas high in introduced alleles (Fitzpatrick and Shaffer 2007). It is believed that A. mavortium prefers perennial ponds because they can extend their larval period to the point of even becoming paedomorphic adults and reproducing while still in the larval form (Fitzpatrick and Shaffer 2007). This life history choice allows them to become sexually mature earlier, breed earlier, and produce larger clutches which gives them an advantage in these habitats and thus maintains a larger introduced population in these ponds (Fitzpatrick and Shaffer 2007).

This situation brings up questions over how to best conserve A. californiense which is currently threatened as well as what is best for the rest of the ecosystem which also includes other endangered species. First, conservationists need to determine which individuals need to be protected and which, if any need to be eradicated (Fitzpatrick and Shaffer 2007). One option would be to preserve the pure natives and eradicate the introduced and hybrid individuals but this only makes sense if the natives can sustain population numbers without the benefits of hybrid vigor. Another would be to protect both the natives and the hybrids, and allow hybrid vigor to increase populations, even if this will destroy the genetic purity of A. californiense. But another aspect to conservation is how it will affect the rest of the system and hybrids may have different effects than natives. It is likely that hybrids, if they possess foraging behaviors of A. mavortium, will be voracious predators of anuran tadpoles and aquatic invertebrates (Fitzpatrick and Shaffer 2007). This could transfer conservation concerns to the prey species because there are endangered species that share ranges with the hybrids and could be additionally harmed by this conservation option (Fitzpatrick and Shaffer 2007). While this threat is yet unproven, species such as the California red-legged frog, Rana draytonii, the western spadefoot, Spea hammondii, and the vernal pool fairy shrimp, Branchinecta lynchi, all share ranges with tiger salamanders and could be harmed (Fitzpatrick and Shaffer 2007).  Conservation of salamanders is already difficult, and hybridization makes it even more so by presenting a multi-faceted problem with many variables to consider.

Chytrid Fungus Poses Threats to Amphibian Species Including Salamanders:

                Another threat to salamander diversity and conservation comes from the fungal species Batrachochytrium dendrobatidis (Bd) which causes a skin disease called chytridiomycosis (Lam et al. 2011). This disease affects different species differently; some species succumb to massive die-offs while others are able to survive with sub-lethal populations on their skin (Lam et al. 2011). Death is likely a result of osmotic and ionic imbalances which then cause cardiac distress and eventually failure (Lam et al. 2011). Bd has two life stages; the first is the infectious zoospore that attacks the keratinized jaw sheaths and tooth rows of larval amphibians (Lam et al. 2011). This second stage is the stationary zoosporangium that focuses on proliferating and spreading the disease (Lam et al. 2011). While the most dramatic mortality events have been recorded in frogs, salamanders are also suffering from the effects of this disease. Some current research is attempting to determine the effects of Bd on fully aquatic salamanders. Chatfield et al. (2012) conducted surveys to first determine the presence of Bd in fully aquatic species of the southeastern US. They captured the salamanders using dipnets and minnow traps and focused their efforts during March-June which was found to be the peak prevalence time (Chatfield et al. 2012). Once they captured a salamander they would use a cotton-tipped swab on its skin to swab multiple times on the dorsum, vent and each foot(Chatfield et al. 2012).Then they used PCR assays to determine whether or not Bd DNA was present on the skin of the captured individuals(Chatfield et al. 2012). 

                Their results (Figure 7) did not show any clear geographic distribution or any difference in prevalence between states despite being at similar latitudes and likely similar climates (Chatfield et al. 2012). 

Figure 7: No clear geographic pattern found in prevalence other than that Bd was found in all populations (Chatfield et al. 2012).

But they did find a significant time period where infections peaked (Chatfield et al. 2012). Figure 8 clearly shows a peak period during March to May during which infection rates were the highest (Chatfield et al. 2012). The decline during June (Figure 8) corresponds with a similar decline in semi-aquatic species found in the same region (Chatfield et al. 2012). 

Figure 8: Peak prevalence from March to May with declines in June that lasted throughout the summer months (Chatfield et al. 2012).

The most disturbing results were that Bd was present in all four of the genera they were sampling for (Figure 9) (Chatfield et al. 2012). 

Figure 9: Not only was Bd found in all localities surveyed, but it was also found in all species caught (Chatfield et al. 2012).

One reason for this could be that the aquatic habitats that these genera are present in provide more stable thermal conditions that favor year-round infection of salamanders (Chatfield et al. 2012). If this is true these fully aquatic species may be serving as a reservoir to maintain Bd populations, allowing them to infect semi-aquatic salamanders that become aquatic during their breeding seasons (Chatfield et al. 2012). One reason this could be possible is that both Cryptobranchus alleganiensis and Necturus maculosus live at temperatures near the lower end of Bd’s optimum growth range, while Amphiuma tridactylum is at higher temperature but even these are within the range for Bd (Chatfield et al. 2012). Another is that Amphiuma species readily use crayfish burrows as retreat sites which double as suitable microhabitats for Bd growth (Chatfield et al. 2012). This growth may be augmented when the Amphiuma aestivate and remain inactive, allowing for increased Bd growth (Chatfield et al. 2012). Siren and Pseudobranchus display similar behavior that may increase Bd concentrations; they burrow into substrate to form a desiccation-resistant cocoon that allows them to survive in seasonal wetlands or during droughts (Chatfield et al. 2012).

                It is not clear yet how Bd is affecting fully aquatic salamanders but it is likely that it adds to the effects of other stressors to reduce survival and reproduction (Chatfield et al. 2012). The main effect on these species seems to be that they are able to sustain large populations of Bd and thus act as vectors for transmission to semi-aquatic salamanders (Chatfield et al. 2012). One reason the aquatic species may have avoided the massive die-offs found in other species is that they have large mucus secretions that may contain anti-fungal components that combat the effects of Bd (Chatfield et al. 2012). These characteristics are found in many other amphibian species and provide hope that perhaps through genetic engineering these antifungal characteristics from skin bacteria could fight the Bd epidemic (Chatfield et al. 2012). These bacteria are symbionts that live on amphibian skin and produce antifungal metabolites to either impede fungal growth or directly kill the fungus (Lam et al. 2011). Research into bio-augmentation of these symbiotic bacteria could significantly decrease this disease and help conserve both salamanders as well as other amphibian species(Lam et al. 2011). 

Difficulties in Adapting to Changing Climates and Global Warming:

                Global warming is causing major changes in global and local climates which is affecting salamander species because they are often only adapted to live within a certain range of temperatures (Bernardo and Spotila 2006). Species can respond one of three ways to climate change; they can either adapt behaviorally or evolutionarily, shift their ranges, or fail to adapt or move and die (Bernardo and Spotila 2006). Problems associated with ranges shifts were previously discussed in the limited range section, and there are separate problems for species facing climate change that lack the ability to adapt. It is estimated that climate change will lead to extinction for about one quarter of species because they will not be able to change ranges or adapt (Bernardo and Spotila 2006). Most often acute or chronic warming-induced stress is the selection factor and should species not be able to adapt to tolerate higher temperatures they will go extinct (Bernardo and Spotila 2006).Bernardo and Spotila (2006) studied thermal tolerances in salamander species and their effects as stressors on the species. In this way they were able to determine how changing climates will affect salamander species that are adapted to cool mountain climates (Bernardo and Spotila 2006).  These species are of special concern because it is likely that they lack the physiological capacity to either tolerate the higher stress levels, or to disperse away from their historic ranges which are now unsuitable (Bernardo and Spotila 2006).This makes these specialized species especially vulnerable to declines which then leads to genetic erosion and likely extinction (Bernardo and Spotila 2006).

                Using clinally varied populations of Desmognathus salamanders, Bernardo and Spotila (2006) were able to study the effects of temperature tolerance within species that have extensively different thermal and prey resource environments. They measured the resting metabolic rates as a function of temperature of low and middle elevation salamanders within their natural thermal gradients (Bernardo and Spotila 2006). They captured specimens of D. carolinensis, and D. ocoee which they then housed without food for at least 72 hours in 250ml plastic boxes with moist moss (Bernardo and Spotila 2006). These boxes were then stored in a room that experienced natural photo- and thermoperiods, and temperatures equivalent to field conditions (Bernardo and Spotila 2006). Results showed that the salamanders had highly conserved metabolic responses to a range of field temperatures (Bernardo and Spotila 2006). Metabolic rates increased over the thermal range in all the populations, and showed significant metabolic depression between 15 and 20 degrees Celsius (Figure 10) which is nested well within the range of normal summer temperatures experienced by these species (Bernardo and Spotila 2006). Metabolic depression varied inversely with elevation causing the lowest elevation salamanders to suffer 42.8-55.5% depressions (Figure 10) (Bernardo and Spotila 2006). 

Figure 10: Metabolic depression increases with decreasing latitudes so low elevation salamanders are experiencing extremely high levels which means they are reaching their thermal tolerance limit (Bernardo and Spotila 2006).

These levels are the highest reported for any amphibian other than aestivating sirenid salamanders which suggests that these salamanders are already living near the limit of their tolerance (Bernardo and Spotila 2006). This would mean that they are unlikely to be able to leave their range to disperse throughout the surrounding warmer, drier valleys (Bernardo and Spotila 2006).

                Range restrictions will continue with climate change, moving the lower tolerance limits upslope and causing even more range contraction (Bernardo and Spotila 2006). These contractions will lead to the same problems as species found in small ranges. They will have to deal with population genetic changes such as decreases in variability due to small effective population sizes and reduced emigration and immigration because of limited dispersal capabilities (Bernardo and Spotila 2006). Another problem in this location would be both competitors and predators moving upslope that are more xeric-adapted (Bernardo and Spotila 2006). These species could then cause further population declines in the native species that are not able to disperse away into more suitable habitats (Bernardo and Spotila 2006). Thus range contraction will intensify genetic fragmentation in these species, who often also are dealing with other forms of habitat modification (Bernardo and Spotila 2006).  Cool, montane systems are significantly structured by strong interspecific interactions and range shifts may have negative effects on these interactions and be detrimental to the stability of montane systems and the specialist species that inhabit them (Bernardo and Spotila 2006). Climate change will also cause these systems to become warmer and drier which could additionally affect the prey resources of these species yet again causing declines (Bernardo and Spotila 2006).

                Their findings demonstrate how not all species will be able to adapt to climate change. The metabolic depression that these species experience even now within their natural ranges will preclude them from being able to adapt to warming (Bernardo and Spotila 2006). As a result of climate change these more specialized species face genetic erosion and likely extinction (Bernardo and Spotila 2006). This trend is likely to hold true globally and organisms that are adapted to cool environments are more likely to have higher extinction rates (Bernardo and Spotila 2006). Globally, this fact will mean that extinction estimates may be underestimates because they do not account for how many cool habitat species face extinction (Bernardo and Spotila 2006). This is because montane systems tend to have higher biodiversity and more specialists (Bernardo and Spotila 2006). Thus, global warming and climate change have the potential to hugely endanger specialist species in addition to compounding the effects of other threats (Bernardo and Spotila 2006). 

What is Currently Being Done to Save Salamander Diversity:

                The threats facing salamander species are gaining attention and thus more research is being focused on how to counteract these threats. One common way that humans work to conserve biodiversity in the face of human projects is to mitigate any habitat losses to species that occur as a result of human projects. This idea was proposed by the Fish and Wildlife Coordination Act of 1958 to compensate for the take of a certain number of individuals of an endangered species in cases when the take cannot be prevented (Searcy and Shaffer 2008). Mitigation efforts attempt to ensure that there is no reduction in species survival, but the difficulty comes in how to determine how much mitigation will counteract the effects of the take (Searcy and Shaffer 2008). As of 2008, the way that most mitigation efforts were calculated was on a basis of credits which are assigned to the affected habitat which then the mitigation will replace elsewhere (Searcy and Shaffer 2008). The parties responsible for the habitat modification will then either protect or enhance other habitat to equal the same number of credits as their take (Searcy and Shaffer 2008). The credit ratios range from 1:1 to 2 or 3:1 in which the party will be mitigating with habitat that contains more mitigation credits than the destroyed habitat (Searcy and Shaffer 2008). Unfortunately these mitigation efforts have not been successful at in-kind habitat replacement based on vegetation criteria, to the point that of a survey of 45 mitigation wetlands only 49% were successful (Searcy and Shaffer 2008).  In another survey 30 wetlands were scored from 0-10 and the average score was 4.66 which means that these efforts are not having the effects they should (Searcy and Shaffer 2008). Part of the problem for these unsuccessful mitigation efforts is that the credits are assigned as a simple constant conversion ratio that equates a certain habitat area with a certain number of credits which then means the same amount of area would need to be enhanced elsewhere for a 1:1 credit ratio (Searcy and Shaffer 2008). The way of calculating credits ignores the quality of the habitats and how they function for the species that use them (Searcy and Shaffer 2008). However, in practice this is difficult and can lead to problems of uncertainty, potential irreproducibility, and subjective nature (Searcy and Shaffer 2008). Searcy and Shaffer (2008) examine the practicality of mitigation estimates and if they are truly enough to support the affected species.

                It is obvious that a new way of calculating mitigation credits is needed, one that takes into account the biological significance of the habitats in question (Searcy and Shaffer 2008).  Searcy and Shaffer (2008) propose to assess habitat quality based on the product of density, individual survival probability, and mean expectation of future offspring. They tested this new method on populations of California tiger salamanders, Ambystoma californiense, located at Jepson Prairie Reserve in the California Central Valley (Searcy and Shaffer 2008). They chose this system because there has been a lot of recent research on the life history of this species and that its habitat is in danger because the Central Valley faces rapid urbanization which means that mitigation will have a huge result on the future of A. californiense (Searcy and Shaffer 2008). First they split the population into three age groups which are visually identifiable: metamorphs, juveniles, and adults (Searcy and Shaffer 2008). Overall they collected 5582 salamanders, including 608 adults, 1828 juveniles, and 3146 metamorphs, using drift fences and paired pitfall traps at the ends of each line. They collected data for each salamander caught including the distance from the pond which allowed them to calculate the density per distance away from the pond (Searcy and Shaffer 2008). A decreasing trend in salamander density away from the shoreline is demonstrated in Figure 11 (Searcy and Shaffer 2008). 

Figure 11: Salamander density decreases drastically away from the shoreline in all age classes (Searcy and Shaffer 2008).

They were then able to calculate survivorship based on the fact that the salamanders reach maturity around four years where metamorphs had a .08 probability of survival and juveniles had a .37 chance (Searcy and Shaffer 2008). This data was then used to calculate the density distribution of reproductive value compared to distance from the shoreline (Figure 12) (Searcy and Shaffer 2008). 

Figure 12: Corresponding to decreasing salamander density, reproductive values of the habitat decreases away from the shoreline (Searcy and Shaffer 2008).

From here they were able to determine a mitigation ratio which decreases in proportion to the density of salamanders and likewise decreases away from the shoreline (Figure 13) (Searcy and Shaffer 2008). 

Figure 13: Because reproductive habitat value decreases with distance Searcy and Shaffer (2008) suggest that mitigation ration should decrease so that habitat near the shoreline should be worth more credits.

This new ratio did no change the total cost of mitigation but rather redistributed the cost in terms of what portions of the landscape are most used by the salamanders so those that were more important and used more had higher values (Searcy and Shaffer 2008).

                There are multiple benefits to this new way of calculating mitigation credits that actually provide economic incentives for parties who follow this new method (Searcy and Shaffer 2008). The first, as previously mentioned, is that this does not increase mitigation costs but just changes the way that credits are assigned (Searcy and Shaffer 2008). The second is that by making certain areas more valuable that others (for example the shoreline is much more valuable than further away from it where there are fewer salamanders) developers may decide to avoid these areas completely in favor of less valuable parcels that would require less mitigation (Searcy and Shaffer 2008). This gives developers incentives to protect areas such as breeding ponds which are ecologically significant for these endangered species (Searcy and Shaffer 2008). Essentially this method assumes that not all habitats are created equally for endangered species and thus the biological uses must be taken into account when trading habitats in one area for ones elsewhere. In this manner, the new calculation method is a much better way of estimating habitat replacement and thus the conservation of endangered species including salamanders.

Conclusion:

While my blog only looks at the research that has gone into salamanders in the United States, there is plenty additional research happening on salamander species worldwide. This plethora of research indicates that salamanders are no long the same mysterious species they once were. A shift has occurred and now increased curiosity in salamander species is causing more researchers to devote their efforts to understanding the threats to salamanders. The pursuit of knowledge will likely be what saves these species because now that humans are more aware of the plight of salamanders, they are trying to make a difference. As long as humans continue to remain aware of these issues, salamanders have a much higher chance of survival. My aim for this blog was to continue spread this awareness and provide a background into the threats to salamanders in the hopes that more and more people will become interested in these issues, and invested in the survival of these amazing organisms.

References

Bayer, C.S.O., Sackman, A.M., Bezold, K., Cabe, P.R., and Marsh, D.M. 2012. Conservation genetics of an endemic mountaintop salamander with an extremely limited range. Conservation Genetics 13: 443-454.

Bernardo, J. and Spotila, J.R. 2006. Physiological constraints on organismal response to global warming: mechanistic insights from clinally varying populations and implication for assessing endangerment. Biology Letters 2: 135-139.

Chatfield, M.W.H., Moler, P. and Richards-Zawacki, C.L. 2012. The amphibian chytrid fungus, Batrachochytrium dendrobatidis, in fully aquatic salamanders from southeastern North America. PLOS ONE 7(9): 1-5.

"Declining Amphibian Populations Task Force (DAPTF)." Information Center for the Environment (ICE). N.p., n.d. Web. 29 Apr. 2013.

Fitzpatrick, B.M., and H.B. Shaffer. 2007. Hybrid vigor between native and introduced salamanders raises new challenges for conservation. PNAS 104(40): 15793-15798.

Graham, S.P., Beamer, D., and Lamb, T. 2012. Good news at last: Conservation status of the seepage salamander (Desmognathus aeneus). Herpetological Conservation and Biology 7(3): 339-348.

Lam, B.A., D.B. Walton, and R.N. Harris. 2011. Motile zoospores of Batrachochytrium dendrobatidis move away from antifungal metabolites produced by amphibian skin bacteria. Ecohealth 8:36-45.

Peterman, W.E., Crawford, J.A. and Semlitsch, R. D. 2011. Effects of even-aged timber harvest on stream salamanders: Support for the evacuation hypothesis. Forest Ecology and Management 262: 2344-2353.

Searcy, C.A. and Shaffer, H.B. 2008. Calculating biologically accurate mitigation credits: Insights from the California tiger salamander. Conservation Biology 22(4): 997-1005.