Low impact of chytridiomycosis on frog recruitment enables persistence in refuges despite high adult mortality
Introduction
Emerging infectious diseases are increasingly threatening wildlife, and have been implicated in recent high profile die-offs in bats, bees and corals (Daszak et al., 2000, Fisher et al., 2012). The amphibian disease, chytridiomycosis, caused by the fungal skin pathogen B. dendrobatidis (hereafter Bd) has been linked to amphibian declines in Europe, Australia and the Americas (Berger et al., 1998, Lips et al., 2006, Skerratt et al., 2007). Chytridiomycosis has caused the likely extinction of 113 species (Skerratt et al., 2007). Many additional species have experienced declines in abundance and distribution and while some species re-expand (Scheele et al., 2014a), many remain highly restricted in small, remnant populations with endemic Bd infection (Briggs et al., 2010, Puschendorf et al., 2011, Phillott et al., 2013). Understanding how remnant populations persist can assist in their conservation and help develop effective response strategies for other chytridiomycosis-threatened species (Scheele et al., 2014b).
Remnant populations might persist through a range of mechanisms including; unfavourable environmental conditions that limit Bd growth (Heard et al., 2014), predatory microorganisms that consume infectious zoospores (Schmeller et al., 2013), genetic based resistance (Savage and Zamudio, 2011) and changes in pathogen virulence or evolved host-pathogen interactions (Altizer et al., 2003). Additionally, some species appear to persist despite high Bd prevalence and low adult survival (Muths et al., 2011, Phillott et al., 2013). This suggests that compensatory recruitment may be important (Muths et al., 2011, Tobler et al., 2012, Phillott et al., 2013). Consistent with this, simulations indicate that low tadpole infection and subsequent high juvenile frog survivorship can provide a buffer against adult mortality in populations challenged by chytridiomycosis (Louca et al., 2014). However, empirical research investigating the impact of Bd on both adult survivorship and recruitment potential in remnant populations remains limited. Given the hypothesized importance of high annual recruitment for buffering populations against Bd-induced adult mortality, there is an urgent need to determine the effects of the pathogen on early life history stages and possible mechanisms that may facilitate high recruitment in populations persisting with Bd.
Warm environmental conditions may provide a refuge for tadpoles and juveniles to clear Bd infection in some species, facilitating high recruitment and buffering populations against Bd-induced adult mortality. Temperature is well known to limit Bd infection in adults, with exposure to warm conditions in terrestrial (Puschendorf et al., 2011, Rowley and Alford, 2013) and aquatic environments (Forrest and Schlaepfer, 2011, Heard et al., 2014) providing protection against the pathogen. In vitro, Bd growth and survival is reduced at temperatures ⩾28 °C (Piotrowski et al., 2004) and exposure to temperatures of between 27 °C and 37 °C has effectively been used to clear infection in a variety of amphibian species in captivity (Chatfield and Richards-Zawacki, 2011, Baitchman and Pessier, 2013). However, whether warm environmental conditions provide protection from Bd for larval stages in the wild remains poorly studied despite suggestions such mechanisms could contribute to population persistence (see Doddington et al., 2013).
While maintaining recruitment potential depends on low Bd impacts during the vulnerable tadpole-juvenile period, survival through this stage is of little value if individuals subsequently succumb to disease prior to breeding. Thus, it is important to understand the timing of disease impact in adults. Many populations with endemic Bd exhibit large seasonal variation in disease frequency, with periods of high prevalence in adults during the breeding season (Kinney et al., 2011) and under favourable climatic conditions (Phillott et al., 2013). For example, Bd prevalence can rise dramatically when pond breeding amphibians enter their breeding habitat and are exposed to water-borne zoospores (Fisher et al., 2009, Kinney et al., 2011). The timing of this increase relative to breeding is crucial: if a substantial proportion of adults are able to breed prior to large increases in disease prevalence and intensity, the pathogen’s impact on adult reproductive potential in that season may be limited.
In this study we asked how remnant populations of the endangered frog L. v. alpina (alpine tree frog) coexist with endemic Bd infection. We addressed this question using three lines of evidence: (1) disease impact on adult survival and changes in disease prevalence during the breeding season, (2) survival to the juvenile stage disease-free as a measure of recruitment potential, and (3) the role of environmental refugia in protecting tadpoles from infection. L. v. alpina has experienced major declines attributed to chytridiomycosis (Osborne et al., 1999, Osborne and Hunter, 2003) and is highly susceptible to Bd infection (six of 277 adults exposed to Bd under laboratory conditions survived, S. Cashins, James Cook University, unpublished results). However, some remnant populations persist and appear to be relatively stable (Osborne et al., 1999); despite high Bd prevalence. Given high susceptibility under laboratory conditions, but observed persistence with Bd in a small number of locations, we hypothesized that populations maintain sufficient recruitment potential to facilitate persistence despite ongoing Bd-induced adult mortality.
By examining both the impact of chytridiomycosis on adult survival and recruitment potential, our study provides important new insights into mechanisms of population coexistence with disease. Our work has implications for the management of species threatened by chytridiomycosis; demonstrating that there is substantial potential for novel strategies to focus on increasing recruitment potential to prevent further extinctions.
Section snippets
Study area and species
Sampling was replicated across three independent L. v. alpina (alpine tree frog) populations (Kiandra, Eucumbene and Three Mile) located in Kosciuszko National Park (35°34′18″S, 148°32′27″E) in south-eastern Australia between August 2011 and March 2013. Population sizes are relatively small and populations likely consist of less than 150 breeding adults (D.H., unpublished results). Populations are separated by distances of between six and 30 km and range in elevation from 1380 to 1475 m. The
Adult survival
Based on the population age structure data, the annual adult male survival rate was very low (0.04). In both 2011 and 2012, breeding aggregations were dominated by a two year old adult male cohort. This result was consistent across all three populations. In 2011, 10 of 88 sexually mature males were one year old, 77 two years old and only one was three years old. In 2012, 78 of 81 sexually mature males were two years old and three were three years old. The small number of three year old males in
Discussion
Determining how remnant populations of chytridiomycosis-threatened amphibians coexist with Bd may prevent further extinctions. By focusing on both the impact of Bd on adult survival and recruitment potential, our study provides crucial new insights into how populations persist with endemic Bd. We show that adult survivorship between the 2011 and 2012 breeding seasons was very low, likely related to high levels of Bd-induced mortality. However, despite nearly all adults dying during their first
Conclusions
B. dendrobatidis has devastated amphibians globally (Skerratt et al., 2007, Fisher et al., 2009) and is now endemic in remnant populations of many declining species (Briggs et al., 2010, Hunter et al., 2010, Puschendorf et al., 2011). Understanding how remnant populations persist with chytridiomycosis is central to their conservation and for guiding the restoration of other species that are largely extirpated from the wild but are retained in captivity (Scheele et al., 2014b). Our study showed
Acknowledgements
Funding was provided by a Taronga Zoo Field Conservation Grant. In kind field support was provided by the NSW Office of Environment and Heritage. Field and laboratory assistance was provided by C. Scheele, S. Kearney, C. Portway and R. Webb. Research was completed under NSW National Parks and Wildlife Scientific Licences SL100411 and SL100436 and ethics clearances A2011/51 and A041025/02.
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