Pathogen Characteristics

In general, the outcomes of epizootics in amphibian communities strongly depend on the respective pathogen strain (Brunner et al. 2015; Berger et al. 2016). Different pathogens are highly variable in their virulence (commonly measured as the time to the host's death; Berger et al. 2005; Brunner and Collins 2009; Farrer et al. 2011; Echaubard et al. 2014; Morrison et al. 2014), and recent studies have identified the circulation of highly recombinant pathogen lineages in wild populations (Farrer et al. 2011; Claytor et al. 2017; O'Hanlon et al. 2018; Vilaça et al. 2019). The origins of the different Bd lineages, as well as Bsal, are in northeast Asia, where they infect native amphibians without causing severe epizootics, and most individuals sustain sublethal infections (Martel et al. 2014; O'Hanlon et al. 2018). This is likely explained by a coevolutionary history of the hosts and the pathogen, further supported by the detection of the fungi in museum specimens at least 100 yr old (Martel et al. 2014; Lips 2016). The geographical origin of ranaviruses is not clear, but phylogenetic analyses suggest that ranaviruses have been a natural part of amphibian host–pathogen systems in Asia for an extended period of time, and have been introduced into North America within the last 100 yr (Vilaça et al. 2019).

However, only half of the literature determined pathogen identity (e.g., Bd lineage or Ranavirus strain). Given the advancements in sequencing techniques and significant decrease in associated time and costs, we suggest routinely conducting isolate identification and providing sequences on GenBank. This would drastically increase the comparability among studies and provide a greater insight into the identity and diversity of pathogen strains circulating in wild and captive amphibian populations. Furthermore, this knowledge would greatly advance management and conservation strategies.

Other key factors for the outcomes of an epizootic in an amphibian population are the pathogen-specific replication and transmission modes (Kriger and Hero 2007; Gray et al. 2009). Bd infection occurs predominately through motile free-swimming zoospores in aquatic habitats, shed by infected individuals, but also through direct transmission (Rowley and Alford 2007; Chestnut et al. 2014). Similar routes of transmission are expected for the salamander fungus Bsal (Gray et al. 2015; Schmidt et al. 2017). Ranaviruses can be transmitted directly through contact with infectious individuals, but also through virions in water and on environmental substrates (Brunner et al. 2007; Cunningham et al. 2007; Robert et al. 2011; Brenes et al. 2014). Another plausible route of transmission, in particular for larval amphibians, is cannibalism (Brunner et al. 2015). This transmission route has been demonstrated in anurans (Pearman et al. 2004; Harp and Petranka 2006) and salamanders (Brunner et al. 2007). Vector-borne transmission can explain the spread of amphibian pathogens across the landscape, in particular over longer distances. Bd zoospores have been found on the feet and feathers of waterfowl (Garmyn et al. 2012; Burrowes and De la Riva 2017). A similar mode of landscape-level transmission would be plausible to expect for ranaviruses; however, there is no peer-reviewed literature testing this hypothesis. Interestingly, Bd as well as ranaviruses were found in mosquitos at sites with infected turtles (Ranavirus, Kimble et al. 2015; Bd, Gould et al. 2019). In addition, Bd has been found on lizard feet (Kilburn et al. 2011). For ranaviruses as a multiclass vertebrate pathogen there is a wide range of potential vector species including fish, reptiles, and amphibians (Duffus et al. 2015).

Host Characteristics

In general, amphibians exhibit high interspecific variation in susceptibility to pathogen infections (Schock et al. 2008; Hoverman et al. 2011; Searle et al. 2011; Ohmer et al. 2013). Therefore, host-identity can be a driving factor in the epidemiology of emerging pathogens in amphibian communities (e.g., chytrid in various North American amphibian species, Gahl et al. 2011; Ranavirus in Californian amphibian communities, Tornabene et al. 2018; Ranaviruses in amphibian communities in the boreal forest of northwestern Canada, Bienentreu 2019). Because of varying susceptibilities, some species experience devastating mortality events, whereas others sustain sublethal infections and potentially act as a reservoir for pathogens (Reeder et al. 2012; Scheele et al. 2016; Bacigalupe et al. 2017; Bienentreu 2019). In Australian amphibian communities, a high abundance of Common Eastern Froglets (Crinia signifera), a reservoir for Bd, amplified prevalence in sympatric Northern Corroboree Frogs (Pseudophryne pengilleyi, Scheele et al. 2017). A similar pattern has been found by Bienentreu (2019), where the presence of Canadian Toads (Anaxyrus hemiophrys) in a community of up to three hosts drastically increased ranavirus loads in Wood Frogs (Lithobates sylvaticus, or Rana sylvatica) and Boreal Chorus Frogs (Pseudacris maculata). With reservoir hosts potentially sustaining sublethal infections and acting as superspreaders for a pathogen, it becomes critical to consider host identity in epidemiological studies.

Additionally, the behavior and ecology of host species can have direct and indirect effects on pathogen dynamics in communities, and vice versa (Lips et al. 2003; Han et al. 2008; Daly and Johnson 2011; Haislip et al. 2012). Contact frequency with other individuals, as well as habitat preferences, can affect transmission (e.g., ATV in Tiger Salamanders [Ambystoma tigrinum], Brunner et al. 2004; Bd in Boreal Toads [Anaxyrus boreas], Hossack et al. 2013). For example, breeding aggregations can facilitate ranavirus transmission (Miller et al. 2011), and chytrid infection probability can be directly linked to thermal preferences of the host species (Rowley and Alford 2013).

Infected individuals (in particular, sublethally infected) may exhibit behavioral changes due to changes in morphology or physiology, subsequently altering transmission dynamics in a host community (Han et al. 2008). Such mechanisms were observed in relation to antipredator behavior, where Wood Frog tadpoles exposed to predator cues exhibited reduced Bd loads, likely due to a stress-induced upregulation of the immune system (Groner and Relyea 2015). When exposed to visual and chemical predator cues, Northern Leopard Frog (Lithobates pipiens, or Rana pipiens) tadpoles infected with Bd exhibited behavioral changes that decreased their risk of being preyed upon, as compared with uninfected individuals (Parris et al. 2006). In contrast, predator cues seemingly do not interact with disease dynamics in the amphibian–ranavirus system. Four larval North American anuran species did not experience elevated mortality or viral loads when exposed to predator cues (Haislip et al. 2012). Another study could not find any relationship between predator cues, resource availability, and ranavirus susceptibility of Wood Frog tadpoles (Reeve et al. 2013). It is also noteworthy that Tiger Salamander larvae infected with ATV, a Ranavirus species commonly found in salamanders, showed drastically increased mortality when exposed to predator cues (Kerby et al. 2011). For echinostomes, however, research showed that increased resource availability can be indirectly linked to the severity of infection in larval anurans by influencing host body size (Marino 2016). Therefore, infection-induced changes in antipredator behavior become relevant when investigating predator–prey systems, where both parts are susceptible to the pathogen (e.g., Bd in frog–salamander system, Han et al. 2011). The vast majority of the literature (91%) included host-specific infection rates, susceptibility, or mortality. However, although 40% of the literature reported pathogen-induced changes in physiology, morphology, or at the cellular level, only 11% investigated effects on growth or development, another 6% investigated behavioral traits and specific host ecology, and only 4% focused on the effects of resource type and availability, as well as predation.

Intraspecific variation in susceptibility plays an important role in pathogen transmission and infection outcome (Echaubard et al. 2010, 2016; Tobler and Schmidt 2010). Such variation typically reflects phylogeny and life history of the host species and their related innate and adaptive immune responses (Grayfer et al. 2015a; Grogan et al. 2018), but also the number and timing of exposures to the pathogen (Hanlon and Parris 2014; Echaubard et al. 2016; Kirschman et al. 2018). The different amphibian life stages present distinct morphologies and physiologies, and their susceptibility to infection changes throughout development (Haislip et al. 2011; Warne et al. 2011). Premetamorphic and metamorphic stages (tadpoles and metamorphs) often exhibit reduced and delayed immune responses to infection in relation to postmetamorphic (juveniles and adults) individuals (Rollins-Smith 1998; Andino et al. 2012; Grogan et al. 2018). Therefore, late-stage tadpoles and metamorphs often show a high susceptibility and commonly succumb to ranavirus infection (Hoverman et al. 2011; Reeve et al. 2013; Grayfer et al. 2014), whereas postmetamorphic individuals can harbor considerably higher pathogen loads than premetamorphic individuals, sustaining sublethal infections without experiencing a terminal outcome (Landsberg et al. 2013; Grayfer et al. 2015b; An and Waldman 2016). The respective amphibian life stage at the time of exposure to a pathogen plays a significant role in infection outcome (Johnson et al. 2011; Echaubard et al. 2016; Kirschman et al. 2018) and despite the differences based on life history stage, only 25% of the literature assessed both pre- and postmetamorphic phases together. Overall, the postmetamorphic stages are more studied, with 63% of the literature compared with only 25% for premetamorphic stages, and only 14% invetigating both stages (Fig. 2B). In studies of chytrid fungi, many studies focus on postmetamorphic host stages (72%), with only 15% of the literature investigating premetamorphic stages and 13% combining both stages (Fig. 3B).

In contrast, in the ranavirus literature, pre- and postmetamorphic stages are balanced (37% each), with 26% of the studies having both life stages combined (Fig. 3D). For helminths, approximately half of the literature (49%) investigated effects on premetamorphic individuals, in comparison with 43% for postmetamorphic individuals, and 8% for both stages (Fig. 3F). For other pathogens such as bacteria and water molds, the majority of the literature (94%) investigated infections in postmetamorphic individuals, and no study included both life stages.

Another important factor in the outcome of pathogen emergence is the phylogeny of the affected host population (Hoverman et al. 2011; Echaubard et al. 2014; Bradley et al. 2015; Bacigalupe et al. 2017). A repeated emergence of a pathogen in a host population will lead to selection toward more resistant genotypes, overall increasing the resilience of the population against new or reintroduced pathogen lineages (Pearman and Garner 2005; Bacigalupe et al. 2017; Puschendorf et al. 2019). The rediscovery of relict amphibian populations of species that were believed to be extirpated or extinct due to epizootics and devastating die-offs (e.g., Armored Mist Frogs [Litoria lorica] and Waterfall Frogs [Litoria nannotis] in Australia, Puschendorf et al. 2011; amphibian communities in the Costa Rican highlands, García-Rodríguez et al. 2012; Harlequin Toads [Atelopus varius] in Costa Rica, González-Maya et al. 2013) may further indicate the coevolutionary history of the host–pathogen system. In fact, pathogens are often present in the host community, but positive individuals show no signs of disease, presumably sustaining sublethal infections (Whitfield et al. 2013; Warne et al. 2016; Puschendorf et al. 2019). However, the phylogenetic backgrounds of the hosts have only been investigated in 11% of the literature.

Host Community and Density

The diversity and dynamics of a community of host species can have a significant influence on pathogen dynamics and vice versa (Becker et al. 2014; Han et al. 2015; Bienentreu 2019; Johnson et al. 2019). Pathogen transmission and infection prevalence are influenced by diversity, density, and abundance of suitable host species (and life stages) in the affected system (Miller et al. 2011; Venesky et al. 2014; Brunner et al. 2015; Muths et al. 2020). In the field of disease ecology, so-called dilution and amplification effects received a great deal of attention in recent years, whereby high host diversity dilutes disease risk and low diversity has amplifying effects (Ostfeld and Keesing 2012; Halliday et al. 2017). However, such relationships have been infrequently investigated in amphibian host–pathogen systems, and the sparse literature shows highly context-dependent interactions. For example, dilution effects were linked to specific tadpole feeding behavior, where the Bd zoospore abundance significantly decreased with an increase in filter-feeding Green Treefrog (Hyla cinerea) and Eastern Narrowmouth Toad (Gastrophryne carolinensis) tadpole diversity (Venesky et al. 2014). A more general relationship was shown in wild Californian amphibian communities of up to six hosts, where an increase in community richness decreased the number of competent host species for the trematode R. ondatrae, reducing the overall infection risk (Johnson et al. 2013). Interestingly, the opposite pattern has been observed for Ranavirus in the same system, where an increase in community richness was positively correlated with ranavirus prevalence (Tornabene et al. 2018).

The effects of community composition become relevant when considering human-mediated species translocations, as well as climate change-induced migration of amphibian species to previously unsuitable habitats. Predictive modeling showed that climate-driven amphibian migrations are expected to be more rapid and advancing farther north than most other vertebrates (Araújo et al. 2006; Lawler et al. 2009). Changes in community diversity can thus lead to the introduction of new pathogens, but also change the dynamics of the pathogens present, by an increase or decrease in competent host species. (Schock et al. 2010; Johnson et al. 2013). In addition, the majority (>80%) of amphibian pathogens is capable of infecting multiple host species within a community, and hosts in natural communities are likely to be infected with multiple pathogens (Blaustein et al. 2011; Olori et al. 2018). In addition, pathogen diversity and abundance in amphibian communities can be increased or decreased by human-mediated habitat fragmentation (King et al. 2007, 2008; McKenzie 2007). Yet, less than half of the literature assessed more than one host species (44%) and only 9% reported more than one pathogen. Furthermore, most of these studies did not directly address potential combined effects of multiple pathogens, and simply stated their co-occurrence. Therefore, epidemiological studies, risk assessments, and management strategies should routinely consider potential effects of host community assemblage and pathogen diversity.

Host density can potentially affect disease dynamics, but only 3% of the literature has considered such effects. In the Sierra Nevada system, higher densities of Mountain Yellow-legged Frogs (Rana muscosa) were positively correlated with a higher number of environmental Bd zoospores and increased infection and mortality (Rachowitz and Briggs 2007; Briggs et al. 2010). Experimental trials with Northern Leopard Frog tadpoles showed that higher density had negative effects on the overall fitness (growth and development), and as a result, animals died faster and at a higher rate when exposed to Ranavirus (Echaubard et al. 2010). By contrast, ranavirus epizootics in Wood Frog populations in Connecticut were strongly influenced by tadpole developmental stage and water temperature, rather than by density (Hall et al. 2018). Similarly, in amphibian communities in Missouri, there was no correlation between Bd presence and community density or species density (Strauss and Smith 2013). Overall, these contradicting patterns show our limited knowledge of density effects on amphibian host–pathogen systems and underline the necessity for further research.

Environmental Factors

Environmental factors influence amphibian host–pathogen systems (Kiesecker 2002; Blaustein et al. 2003), further contributing to population declines (Daszak et al. 2003; Kiesecker et al. 2004; Lips et al. 2006). Over the last decade, environmental factors were considered in 41% of the research on amphibian pathogens. In particular, a third of these studies (32%) investigated landscape features such as slope and elevation, habitat characteristics such as vegetation type and canopy cover, and also natural disturbances such as wildfires and severe weather events, which can potentially affect pathogen dynamics in amphibian communities (Becker et al. 2012; Hossack et al. 2013; Muths et al. 2020). These alterations are mediated by changes in temperature, environmental humidity, and light intensity. At amphibian breeding sites in Maine, low canopy cover was positively associated with ranavirus presence and mortality (Gahl and Calhoun 2010), likely due to a higher mean water temperature. The opposite effects have been observed in endangered rain-forest frogs (Litoria rheocola) in Australia, where the reduction of canopy cover drastically decreased the Bd infection risk for the frogs, mediated by an increase in microhabitat mean temperature and related decrease in environmental humidity (Roznik et al. 2015).

In general, climate is considered a major factor driving amphibian disease dynamics (Rohr et al. 2013). In particular, temperature received increased attention and is a factor in 37% of the literature. Temperature is believed to directly influence the dynamics of emerging pathogens such as Bd in wild amphibian populations (Kiesecker et al. 2001; Woodhams et al. 2003; Berger et al. 2004). Chytrid fungi and ranaviruses are extremely temperature dependent, as reflected in their spatiotemporal distribution patterns (Olson et al. 2013; Feldmeier et al. 2016; Hall et al. 2018; Youker-Smith et al. 2018). Chytrid fungi occur mostly in regions with low mean temperatures and Bd-related declines occur predominantly at higher altitudes (Ron 2005; Woodhams and Alford 2005), whereas Bsal-related declines are currently restricted to forested areas in central Europe (Feldmeier et al. 2016; Watts et al. 2019). Chytrid fungi show optimal growth at temperatures between 10°C and 20°C and decreased growth between 25°C and 28°C, with temperatures over 30°C killing the fungus (Piotrowski et al. 2004; Martel et al. 2013; Stevenson et al. 2013). Ranavirus epizootics often occur in mid- to late summer in temperate regions of the Northern Hemisphere when water temperatures surpass 25°C (Brunner et al. 2015). In fact, ranaviruses show optimal growth at temperatures of 28°C but do not replicate over 31°C (Ariel et al. 2009; Nazir et al. 2012). At temperatures below 12°C, infection loads are low and mortality is drastically reduced (Brand et al. 2016).

Environmental substrates and humidity strongly influence the environmental persistence of pathogens (Nazir et al. 2012; Kolby et al. 2015a) and have been investigated in 20% of the literature. In particular, the type of substrate (e.g., inorganic and organic) as well as humidity (e.g., precipitation, water connectivity, substrate moisture) affect the short-term viability of the pathogens (Johnson and Brunner 2014; Kolby et al. 2015b; Munro et al. 2016; Stoler et al. 2016). Therefore, it is unlikely that pathogens show long-term persistence in environmental substrates. Ranaviruses and chytrid fungi may rely therefore on reservoirs, such as sublethally infected individuals, for long-term persistence in amphibian communities (Gray et al. 2009; Blaustein et al. 2011; Bosch et al. 2015; Schmidt et al. 2017; Brannelly et al. 2018). In fact, the amphibian-reservoir hypothesis is strongly supported by reoccurring pathogen emergence in amphibian communities inhabiting semipermanent and ephemeral wetlands (Hunter et al. 2010; Hall et al. 2018; McMillan et al. 2020), as well as after environmental chemical disinfection (Bosch et al. 2015). Furthermore, Bd and ranaviruses can persist in overwintering individuals (Brunner et al. 2004; Narayan et al. 2014).

UVB radiation can also cause an increased vulnerability to infection by impairing growth and development and possibly inducing malformations (Blaustein et al. 2003; Searle et al. 2010). Ambient UVB radiation significantly decreased Bd prevalence in infected larval Common Toads (Bufo bufo) and European Midwife Toads (Alytes obstetricans, Ortiz-Santaliestra et al. 2011), but no such connection was found for ranaviruses (Gahl and Calhoun 2010). However, only 1% of the literature assessing environmental factors associated with amphibian pathogens considered UVB radiation as a potential explanatory variable for epidemiological patterns. Similarly, as little as 4% of the literature has assessed water parameters such as salinity, conductivity, pH, or dissolved oxygen. Elevated salinity significantly reduced pathogen transmission and infection loads and increased survival rates among Bdinfected amphibians in Australia (Stockwell et al. 2015a,b; Clulow et al. 2018). The opposite pattern was observed in a Wood Frog–ranavirus system, whereby an increase in chloride concentrations slightly increased the chances for a mortality event (Hall et al. 2018). Other factors such as pH and conductivity had no significant effects on amphibian host–pathogen systems (e.g., Ranavirus in Wood Frogs in Connecticut, Hall et al. 2018; Bd in ranids in Missouri, Strauss and Smith 2013). However, in a study across the continental United States, Bd loads of swabbed frogs were lower at sites with elevated pH levels (Battaglin et al. 2016).

Overall, there is very little support for general effects of natural water characteristics such as pH, conductivity, and salinity on pathogen dynamics. The literature clearly shows that interactions are highly context dependent. Yet, water characteristics can indirectly influence pathogen dynamics through alteration of the microbiome involved in Bd tolerance (Varela et al. 2018) or by directly affecting hosts' overall body condition and related resilience and immunity, making additional research on these relationships necessary.

Anthropogenic Factors

There is substantial evidence for a direct correlation between anthropogenic factors and epidemiological patterns in amphibian populations (e.g., Schotthoefer et al. 2011; Koprivnikar and Redfern 2012; Tornabene et al. 2018). Sixteen percent of the literature investigated such relationships and found that industrial and infrastructural development has strong effects on host–pathogen dynamics by altering habitats and host community structures (St-Amour et al. 2008; Belasen et al. 2018; Davis et al. 2019).

Highly virulent pathogen lineages from commercial trade and breeding facilities (Farrer et al. 2011; Claytor et al. 2017; Vilaça et al. 2019) and human-mediated species translocation (e.g., use as bait) accompanied by the introduction of new pathogens represents a severe threat to naïve populations, and can lead to fast and devastating mass mortality events (Picco and Collins 2008; Price et al. 2014; Earl et al. 2016). Research showed that commercial frog farms facilitated strain recombination events (Bd, Schloegel et al. 2012; Greenspan et al. 2018; Ranaviruses, Claytor et al. 2017; Vilaça et al. 2019) and can release ecologically relevant amounts of infectious agents in the adjoining ecosystem by discharge of untreated wastewater (e.g., Bd zoospores, Ribeiro et al. 2019). Similar mechanisms of pathogen spread were suggested for the amphibian–ranavirus system (Saucedo et al. 2019). In addition, ranavirus isolates collected from epizootics in commercial frog-breeding facilities (Majji et al. 2006; Miller et al. 2007) showed significantly increased virulence in a variety of different amphibian species in comparison with wild-type isolates (Hoverman et al. 2010, 2011; Peace et al. 2019).

In addition to breeding facilities, the associated trade of the animals contributes significantly to the global spread of amphibian pathogens such as Bd and ranaviruses (Picco and Collins 2008; Kolby et al. 2014). A study investigating the presence of common pathogens in commercially traded amphibians at the Hong Kong International Airport revealed that about 1 in 10 individuals was infected with Bd, and more than half tested positive for ranaviruses (Kolby et al. 2014). High densities during import and export can amplify virulence and transmission among individuals (Pavlin et al. 2009). For example, Tiger Salamanders in the North American fishing bait trade were infected with highly virulent strains of the Ranavirus species ATV (Jancovich et al. 2005; Picco and Collins 2008; Epstein and Storfer 2016). Furthermore, invasive Bullfrogs (Lithobates catesbeianus, or Rana catesbeiana) can act as reservoir hosts and can contribute to the emergence and spread of common amphibian pathogens Bd (Garner et al. 2006; O'Hanlon et al. 2018) and ranaviruses (Schloegel et al. 2009).

Pollution from fertilizer, herbicides, and pesticides may indirectly affect pathogen dynamics in aquatic communities by eutrophication and subsequent increase in host density and prevalence (Johnson et al. 2007). Furthermore, many of the commonly used herbicides and pesticides negatively affected the body condition of amphibians (Egea-Serrano et al. 2012), and therefore can influence disease dynamics in host communities. For example, Northern Leopard Frogs exposed to environmentally relevant doses of the insecticide carbaryl harbored significantly higher trematode loads than unexposed individuals (Pochini and Hoverman 2017a). Interestingly, the authors did not observe any direct effects on infection rates. Another study used larval Wood Frogs to investigate the effects of the insecticides carbaryl and thiamethoxam on ranavirus transmission and susceptibility (Pochini and Hoverman 2017b). The animals experienced increased mortality; however, the effects on pathogen transmission and prevalence were negligible. In contrast, five North American anuran species exposed to mixtures of commonly used insecticides or herbicides did not experience increased mortality rates, but showed alterations of Bd loads (Buck et al. 2015). Another study conducted across the United States showed a positive correlation of Bd prevalence in amphibian hosts and total fungicide concentrations in the environment (Battaglin et al. 2016). Furthermore, Bd zoospore loads in ranid and hylid frogs increased with dissolved organic carbon, nitrogen, and phosphorus (Battaglin et al. 2016). Although their impacts on the overall health of amphibians are known, certain pesticides have been used in disease control and outbreak management. Bd was eliminated at several wetlands in Spain through a combination of artificial draining, environmental chemical disinfection, and ex situ treatment of individuals with fungicides (Bosch et al. 2015).

A major threat to amphibian populations worldwide, anthropogenic habitat fragmentation can also control host–pathogen interactions by modifying host community structure and density (Beasley et al. 2005; King et al. 2007; Greer and Collins 2008; St-Amour et al. 2008; Becker et al. 2016). For example, deforestation can directly influence amphibian community diversity, richness, and density by alteration of microhabitat climate and species interactions, in turn variously increasing or decreasing the infection risk for Bd (Becker et al. 2016). Increased microhabitat temperatures can suppress the growth of the fungus and negatively affect its environmental persistence (Raffel et al. 2010; Becker et al. 2012). Similarly, spatial and temporal variation of ranavirus transmission and prevalence was indirectly influenced by land use pattern in Tiger Salamanders (Greer and Collins 2008). An increase in ranavirus infection was correlated with the increase in contact rates between hosts, induced by anthropogenic modifications of the wetland structure (Greer and Collins 2008). Amphibian populations inhabiting fragmented habitats in urban settings in the UK also showed an increased ranavirus prevalence influenced by various anthropogenic factors (e.g., level of urbanization, use of chemicals, invasive species; North et al. 2015), and a direct relationship between industrial and infrastructural activity and ranavirus prevalence was observed in Green Frogs (Lithobates clamitans, or Rana clamitans) in Canada, yet not with Bd (St-Amour et al. 2008). A large study among amphibian populations across the United States revealed a similar relationship (Battaglin et al. 2016).

Overall, anthropogenic factors play a significant role in amphibian disease ecology by directly and indirectly altering host–pathogen dynamics. This underlines the importance of comprehensive ecological assessments and subsequent implementation of management and mitigation strategies when conducting infrastructural and industrial development. Furthermore, biosafety measures, such as disinfection protocols at recreational and commercial facilities (e.g., marinas and harbors), should be implemented to limit pathogen spread through contaminated equipment. Finally, stricter regulations need to be imposed on commercially bred and traded amphibians to prevent the further spread of pathogens and their establishment into wild populations.