Panzootic amphibian fungi causes catastrophic and ongoing loss of biodiversity

Highly virulent wildlife diseases contribute to the sixth mass extinction of the Earth (1). One of these is chytridiomycosis, which has caused massive dieback of amphibians worldwide (2, 3). Chytridiomycosis is caused by two species of fungi, Batrachochytrium dendrobatidis [Découverten1998([Discoveredin1998([découverten1998([discoveredin1998(4)]and B. salamandrivorans [Découverten2013([Discoveredin2013([découverten2013([discoveredin2013(5)]. Both Batrachochytrium probably from Asia and their recent spread has been facilitated by humans (5, 6). Twenty years after the discovery of chytridiomycosis, important research has led to a better understanding of its epidemiology (2, 3, 7, 8), but there are still significant gaps in knowledge. First, the global extent of species decline associated with chytridiomycosis is unknown.[see([See([voir([see(2, 9) for initial assessments]. Second, although some regional declines are well studied, global spatial and temporal patterns of chytridiomycosis impacts remain poorly quantified. Thirdly, the ecological and biological characteristics have only been examined for a part of the declining species (ten, 11). Finally, after initial declines, it is unclear what proportion of declining species recover, stabilize at lower abundance or continue to decline. Here we present an overall epidemiological analysis of the spatial and temporal extent of amphibian biodiversity loss caused by chytridiomycosis.

We conducted a thorough review of evidence from multiple sources, including the International Union for the Conservation of Nature (IUCN) Red List of Threatened Species (12), peer-reviewed literature and consultations with international experts on amphibians (S1 data). We categorized the species into five categories of decline and severity corresponding to a reduction in abundance. Species declines were attributed to chytridiomycosis based on the diagnosis of infection causing mortality in the wild or, if this was not possible, evidence consistent with the epidemiological characteristics of the disease. Most of the evidence is retrospective as many species declined before the discovery of chytridiomycosis (S1 data).

We report conservatively that chytridiomycosis has contributed to the decline of at least 501 species of amphibians (6.5% of the described amphibian species, Figures 1 and 2). This represents the largest documented loss of biodiversity attributable to a pathogen and B. dendrobatidis among the most destructive invasive species, comparable to rodents (threatening 420 species) and cats (Felis catus) (threatening 430 species) (13). Losses associated with chytridiomycosis are orders of magnitude higher than other large wildlife pathogens, such as white nose syndrome (Pseudogymnoascus destructans) in bats (six species) (14) or West Nile virus (flavivirus sp.) in birds (23 species) (15). Of the 501 decommissioned amphibian species, 90 (18%) are confirmed or presumed extinct in the wild, and another 124 (25%) are experiencing an abundance reduction greater than 90% (Figures 1 and 2). ). Declines for all but one species (Salamander Salamander affected by B. salamandrivorans) were awarded to B. dendrobatidis.

The decreases were proportional to the taxonomic abundance, with 93% of the severe decreases in anurans (they account for 89% of all amphibian species). Within the anurans, there was a marked taxonomic aggregation of declines, with 45% of the severe declines and extinctions occurring in neotropical genera. Atelopus, craugastor, and Telmatobius (Fig. 2) (16). Chytridiomycosis is fatal for caelians (17), but there has been no Caecilian decline due to the disease, although data is limited. The ability to B. dendrobatidis causing significant declines is attributable to the maintenance of its strong pathogenicity (2, 18), wide range of hosts (8), high transmission rate within and between host species (2, 7) and the persistence of host species in reservoirs and the environment (19). For many species, chytridiomycosis is the main factor of decline, as evidenced by massive mortality precipitated in undisturbed environments (2). In other species, chytridiomycosis works in concert with habitat loss, changing climatic conditions, and invasive species to exacerbate species decline (20).

Most amphibian declines occurred in the tropics of Australia, Mesoamerica, and South America (Fig. 1), supporting the hypothesis that B. dendrobatidis spread from Asia to the New World (6). Asia, Africa, Europe and North America recorded significantly lower declines due to chytridiomycosis, despite widespread B. dendrobatidis (8). The relative absence of documented declines may reflect less knowledge of amphibian populations in Asia and Africa (3, 21), the early introduction and eventual coevolution of amphibians and B. dendrobatidis in parts of Africa and the Americas[forexample([Eg([parexemple([eg(22)], the relatively recent emergence of B. dendrobatidis in West and North-East Africa (6), or inappropriate conditions for chytridiomycosis. It is not known whether chytridiomycosis contributed to the widespread decline of amphibians reported in North America and Europe in the 1950s to 1960s (3, 21, 22) or current declines in enigmatic salamanders in eastern North America. Although the number of new declines has now decreased (Fig 3), further declines could occur if B. dendrobatidis or B. salamandrivorans are introduced into new regions, highly virulent lines are introduced into areas where currently less virulent lines are6) and / or changes in the environment modify a previously stable pathogen-host dynamic (3).

Declines associated with chytridiomycosis peaked globally in the 1980s, between one and two decades before disease discovery (Fig. 3 and Table S1), and coincide with the anecdotal recognition of amphibian decline in the 1990s (23). A second, smaller peak appeared in the early 2000s, associated with increased declines in western South America (Fig 3 and Fig S 1). Regionally, temporal patterns of decline are variable (Figure S1). For example, in parts of South America and Australia, declines began in the late 1970s (2, 24), while in other regions, declines started in the 2000s (25). B. dendrobatidis is associated with ongoing declines in 197 species assessed. Decline in progress after transition to enzootic disease dynamics (19) could be motivated by a lack of effective defense of the host, maintaining a strong pathogenicity (18) and presence of B. dendrobatidis in amphibian and non-amphibian reservoirs (7, 19).

We examined the host's biological characteristics and environmental conditions to understand why some species declined more severely than others, using multinomial logistic regression and taking into account the degree of evidence of chytridiomycosis involvement. in the decline of each species (Fig S2 and Table S2). The severity of the decline was greatest for large species, those that were in areas that were still wet, and those that were strongly associated with perennial aquatic habitats. These trends are likely due to favorable environmental conditions for B. dendrobatidis in humid regions (7), because the fungus dies when it is desiccated, as well as the general pattern of increasing time to maturity in large amphibians, which reduces reproductive potential and compensates for chytridiomycosis mortality (26). Decreases were less severe for species with wide geographic and elevation ranges (Fig. 4), probably due to the increased risk that their range includes adverse environmental conditions. B. dendrobatidis (3) and / or information bias, as population extinctions can be more confidently assessed for species with restricted distributions. Our results are consistent with previous studies showing that the risk of chytridiomycosis is associated with the use of aquatic habitat by the host, at large size and at reduced altitude (ten, 11).

Encouragingly, of the 292 surviving species for which demographic trends are known, 60 (20%) showed initial signs of recovery. However, replenishments generally represent small increases in the abundance of individual populations, not complete recovery at the species level. Logistic regression showed that the probability of recovery was lower for species with more recent or severe declines, for larger or nocturnal species, and for species at higher altitudes (Figure S2 and Figure 2). table S3). When recovery predictors are maintained at their mean value, the risk of recovery of a species from severe decline (> 90%) is less than 1 in 10. The low probability of recovery of high altitude species may be related at climatic conditions conducive to survival. fungal persistence, limited connectivity to source populations, and / or longer generation of hosts (26). Some recoveries may be supported by selection for increased host resistance (18), while simultaneous threat management may have facilitated other recoveries (a promising avenue for conservation interventions) (27). Unfortunately, the remaining 232 species showed no signs of recovery.

The unprecedented lethality of a single disease affecting an entire class of vertebrates underscores the threat posed by the spread of pathogens in a globalized world. Global trade has recreated a functional Pangea for infectious diseases in wildlife, with significant impacts on biodiversity (the present study), livestock (28) and human health (29). Effective biosecurity and immediate reduction of wildlife trade are essential to reduce the risk of spread of pathogens. As the attenuation of chytridiomycosis in the wild has not yet been proven (30), new research and intensive monitoring using emerging technologies are needed to identify species recovery mechanisms and develop new mitigation measures for declining species.

Thanks: We thank Mr. Arellano, Mr. Courtois, A. Cunningham, K. Murray, Ron, R. Puschendorf, J. Rowley and V. Vredenburg for their discussions on the decline of amphibians. The comments of two anonymous reviewers greatly improved the manuscript. Funding: B.C.S. and D.B.L. were supported by the Australian National Environmental Science Program. L.B., L.F.S., T.A.K., and B.C.S. were supported by the Australian Research Council (grants FT100100375, LP110200240 and DP120100811), the NSW Environment and Heritage Office and the Taronga Conservation Science Initiative. S.C., W.B., A.M. and F.P. were supported by grants FWO3E001916 and FWO11ZK916N-11ZK918N from the Foundation for Research in Flanders and grant BOF16 / GOA / 024 from the University of Ghent. S.C. has received support from Foundation FWO16 / PDO / 019 of the Research Foundation Flanders. A.A.A. was supported by the Conservation Leadership Program (0621310), the Vicerrectoría de Investigaciones, the University of Pamplona in Colombia and Colciencias (1121-659-44242). T.C. was supported by the Coordination for the Improvement of Higher Education Personnel. CA was supported by the Amazon Conservation Association, the Amphibian Specialist Group, the Disney World Conservation Fund, the Eppley Foundation, the Mohammed Bin Zayed Species Conservation Fund, the NSF, the Foundation Rufford Small Grants and the Swiss National Foundation. I.D.R.R. was supported by the Spanish Government (CGL2014-56160-P). M.C.F. was supported by NERC (NE / K014455 / 1), Leverhulme Trust (RPG-2014-273) and the Morris Animal Foundation (D16ZO-022). S.V.F. was supported by the USFWS Wildlife without Borders (96200-0-G228), the AZA Conservation Endowment Fund (08-836) and the International Fund for Endangered Species (Conservation International). P.F.Á. was supported by a Mexican Research Council Postdoctoral Fellowship (CONACYT, 171465). T.W.J.G. was supported by NERC (NE / N009967 / 1 and NE / K012509 / 1). J.M.G. was supported by San Francisco University of Quito (Collaborative Grants 11164 and 5447). Mr.H was supported by scholarships from the Elsa-Neumann Foundation and DAAD (German Academic Exchange Service). CAM. was supported by the Atkinson Center for a Sustainable Future and the Cornell Center for Vertebrate Genomics. G.P.-O. was supported by DGAPA-UNAM and CONACYT while on sabbatical at the University of Otago, New Zealand. C.L.R.-Z. was supported by the NSF (1660311). S.M.R. was supported by a CONACYT Problemas Nacionales grant (PDCPN 2015-721) and a UC Mexus-Conacy co-operation grant. THAT'S IT. was supported by the Chilean National Fund for Science and Technology (Fondecyt n ° 1181758). L.F.T. was supported by the São Paulo Research Foundation (FAPESP 2016 / 25358-3) and the National Council for Scientific and Technological Development (CNPq 300896 / 2016-6). J.V. was supported by the NSF (DEB-1551488 and IOS-1603808). C.W. has received support from the South African National Research Foundation. Author contributions: B.C.S., F.P., L.B., L.F.S., A.M. and S.C. designed the research. B.C.S. collected data and coordinated data collection. All authors provided ideas and data. S.C. conducted the analysis with the assistance of B.C.S., F.P., A.M., C.N.F., and W.B.B.C.S., F.P., L.B., L.F.S., A.M., C.N.F. and S.C. wrote the paper with input from all authors. Competing interests: The authors do not declare any conflict of interest. Availability of data and materials: All data is available in the manuscript or additional documents.