Humans are altering natural environments at unprecedented rates, with significant negative impacts for animals (MEA, 2005; Pereira et al., 2010). Wetlands in particular harbour highly diverse biological communities and provide multiple ecosystem services, yet are frequently degraded and destroyed, with over 50 % of global wetland surface area lost during the last century (Mitsch and Gosselink, 2007; Zedler and Kercher, 2005). Consequently, declines of wetland-dependent species are some of the greatest recorded (MEA, 2005). Concurrent with these losses, wetlands are being created to harness the ecosystem services they provide, such as those constructed in urban areas to treat storm water, and settlement dams created in mining areas to store and treat processed materials (Hammer, 1989; Odum, 2016).

Although typically not designed to support and conserve wildlife, these artificial wetlands regularly attract animals, as they superficially resemble natural wetlands and contain the cues used by animals when selecting habitats (often termed secondary wetlands; Dolny and Harabis, 2012). Therefore, secondary wetlands may provide critical habitat, enhancing landscape-level connectivity and promoting biodiversity and population persistence (Bendell-Young et al., 2000; Scheffers and Paszkowski, 2013; Thiere et al., 2009). For some organisms, properly managed secondary wetlands may provide habitat superior to that of natural wetlands. For example, Dolny and Harabis (2012) observed more than twice as many dragonfly species in mine subsidence pools than in reference wetlands, suggesting that this was due to enhanced environmental heterogeneity resulting from abiotic succession processes occurring as a direct consequence of mining. Given this potential, there has been considerable interest in simultaneously satisfying industrial needs and promoting wildlife within secondary wetlands (i.e. multi-objective management; Benyamine et al., 2004).

However, since secondary wetlands are often designed to store contaminated water, or are in areas prone to contamination and human interference, animal inhabitants may experience compromised fitness (e.g. reduced survival and reproduction; Dods et al., 2005; Laposata and Dunson, 2000). Therefore, despite being conceptually appealing, multi-objective management may prove problematic if individual fitness within secondary wetlands is impacted, and particularly if individual fitness impacts scale up to affect population persistence.

Quantifying community composition, population sizes and individual fitness will assist management efforts to maintain and enhance the conservation potential of secondary wetlands. In particular, utilising fast, cheap and non-lethal methods of estimating individual fitness is preferable under certain circumstances. Quantifying developmental instability (DI) is a promising method for estimating the fitness of individuals and provides an indirect proxy for habitat quality (Alford et al., 1999; Burghelea et al., 2013; Tracy et al., 1995). Developmental stability occurs when genotypes repeatedly produce the same phenotype under the same environmental conditions during development (Tracy, et al., 1995; Zakharov, 1992), so levels of deviation from stability (i.e. DI) provide information on environmental quality and individual health. DI is calculated by measuring bilateral structures such as the length of limbs and quantifying disparity from symmetry. Studies have shown that DI increases as health decreases, and that amphibian population declines have preceded periods of increasing DI (Alford et al., 2007). Therefore, DI may be a valuable, non-lethal early warning indicator of the impact of poor-quality wetlands.

Determining the suitability of secondary wetlands for currently threatened taxa in particular should be at the forefront of conservation science. Amphibian populations are currently experiencing dramatic declines around the world, with approximately 40 % of species facing the threat of extinction (Monastersky, 2014; Whitfield et al., 2007). Amphibians are often considered particularly sensitive to environmental contamination due to their physiology and biphasic life cycle (Wake and Vredenburg, 2008). Although numerous factors have been implicated in current population declines, such as climate change, disease and pollution, habitat loss clearly stands out as one of the main threats facing amphibians (Cushman, 2006). Consequently, secondary wetlands such those within sand quarries may provide vitally important habitat. Relative to other mining industries which may release materials into wetlands that are highly toxic to wildlife, such as fossil fuel mining, sand quarry wetlands may be more suitable due to the comparably lower environmental impact of this industry (Anderson and Arruda, 2006; Rowe et al., 1996). However, there are few empirical studies investigating how sand quarry wetlands function as habitat for amphibians or indeed any other animal groups (although see Catchpole and Tydeman, 1975; Eversham et al., 1996; Spencer and Griffith, 2012).

Here, I determine how quarry wetlands within the Kables Sands Quarry, Australia, function as habitat for local amphibians, and investigate the factors that may be influencing amphibian populations and communities. To do this, I measured environmental variables and quantified amphibian population density, species richness and levels of developmental instability (DI) within sand quarry wetlands, and compared findings to reference wetlands and previously published studies.

Wetlands within the Kables Sands quarry were inhabited by more frog species than nearby urban wetlands, and the individuals within quarries had lower levels of developmental instability. Of the species present at the quarry, only Crinia signifera were present within reference wetlands, and these frogs had comparable rates of DI with frogs from populations that underwent crashes (Alford, et al., 2007). I also present evidence that there is a negative relationship between the abundance of C. signifera and unsigned asymmetry of their limbs at particular wetlands. Species occurrence and individual fitness proxies may be used to predict population persistence (Aldridge and Boyce, 2007). Therefore, quarry wetlands may provide more suitable habitat conditions than urban wetlands in the longer term.

Importantly, reference wetlands selected for this study do not necessarily represent pristine natural sites, but rather were artificial wetlands in suburban areas. Therefore, there may be additional pressures on these wetlands that quarry wetlands are not exposed to, such as elevated contaminants from storm water run-off, which influence amphibian richness, abundance and fitness (Hale et al., 2015). Despite not being pristine, amphibians frequently inhabit urban wetlands and these provide essential habitat in many urban areas (Brand and Snodgrass 2010; Le Viol et al., 2012). Since these reference wetlands are omnipresent in human-dominated landscapes, they serve as relevant and useful comparators to determine the suitability and ecological importance of quarry wetlands.

Typically, the mean abundance of Crinia signifera is greater at larger wetlands and at those with greater connectivity (Hamer et al., 2012). Although the site with the highest abundance of Crinia signifera – Q4 – was the second smallest quarry wetland by area, importantly, it also had the largest perimeter. Although there was no overall effect of wetland perimeter, given that this species calls from the waters edge, a greater perimeter may permit more individuals to be present and calling at any one time (Antis, 2013). Whilst salinity and pH were significantly higher within reference sites, and pH may have influenced frogs within this study, the levels observed within all surveyed wetlands were well within the range deemed suitable for the majority of amphibian species native to this region (Kearney et al., 2012; Sadinski and Dunson, 1992; Smith et al., 2007).

Our study and others conducted within Australian wetlands may shed light onto the factors driving community- and population-level patterns, and also provide information on how to best manage quarry wetlands. The occurrence of both Limnodynastes peronii and Litoria ewingii correlate with the proportion of vegetation in the emergent and submerged zones (Hamer et al., 2012). Although we found little influence of vegetation, at wetlands with > 80 % cover, these species have a > 0.8 probability of being found, and yet were not heard calling at the reference wetlands that contained high levels of vegetation. In addition, L. peronii occurrence is correlated to the depth of wetland shores, where gently sloping shores are preferred over steep drops (Hamer et al. 2012). The shoreline of one of the reference wetlands, R4, was made up of a concrete wall, largely inappropriate for successful breeding by ground frog species (Parris, 2006). Within the quarry, there was considerable inter- and intra-wetland variability in the steepness of the shoreline, and given other species within the region (e.g. Limnodynastes dumerlii) prefer steep shores. This variability likely enhances year-round biodiversity in quarries. I provide considerable evidence that for sand quarry wetlands, multi-objective management may be highly beneficial for local amphibian populations (Bendell-Young, et al., 2000).

In addition to quantifying trends in species presence and population size within quarry wetlands, effective management and conservation also require knowledge of individual fitness. Studies have shown that DI increases as health decreases, and amphibian population declines have preceded periods of increasing DI (Alford, et al., 2007). I provide three lines of evidence suggesting quarry wetlands are providing a comparatively high-quality resource. Firstly, relative to reference wetlands, frogs within quarry wetlands had significantly lower levels of DI. Secondly, DI rates within quarry wetlands were comparable to the control populations from Alford et al. (2007). Frog populations within these control sites remained stable, whilst populations from impact sites exhibited substantial declines following a period of increasing DI. Finally, there was a significant relationship between levels of DI and the abundance of C. signifera populations. Although this suggests that population persistence and/or growth may be enhanced when DI is lower, quantifying population dynamics was beyond the scope of this study. Furthermore, whilst the incidence of amphibian abnormalities has been shown to be higher in wetlands impacted by other forms of mining (Anderson and Arruda, 2006; Rowe et al., 1996), I identified no abnormalities in any of the frogs captured in this study.

Empirical studies that monitor the relationship amongst population size, community composition, individual fitness and the environmental factors driving these responses are imperative to the long-term persistence of amphibians in human-impacted environments. Given that the quarry wetlands surveyed here provide a seemingly good habitat for amphibians, they – and similar wetlands elsewhere – should be managed to enhance their suitability and attractiveness. Gallagher et al. (2014) suggest ensuring that high-quality wetlands are connected to other wetlands, with their hydroperiod and vegetation managed to promote amphibian survival and reproductive success. Measures to remove and exclude exotic predators, such as mosquitofish and carp, will also greatly enhance amphibian populations and communities, with considerable improvements to individual fitness (Maezono and Miyashita, 2004; Tsunoda et al., 2010). To further promote and enhance amphibian biodiversity within quarries, it is important to provide substantial environmental heterogeneity; a mosaic of both wetland types and terrestrial refuges. Creating a diverse mosaic of wetlands that vary in size and depth will help ensure amphibian breeding at a proportion of quarry wetlands each season, thereby enhancing overall population and community persistence at any one quarry (Rannap et al., 2009). Despite little influence of vegetation on richness or abundance found here, creating wetlands with a minimum of 80 % emergent vegetation will help ensure a high probability of many native species occurring, such as Limnodynastes peronii and Litoria ewingii (Hamer et al., 2012). All species identified here oviposit their eggs within and around aquatic and semi-aquatic vegetation, which in turn provide refuge and protection to tadpoles (Anstis, 2013). Therefore, planting native vegetation around and within these wetlands would greatly increase the probability of colonisation. In quarries with existing wetlands, such as Kables Sands, implementing topographical modifications will ensure sufficient heterogeneous habitat to cater for as many species as possible.

In terms of monitoring, identifying amphibian species presence and population size and quantifying individual fitness by measuring limb lengths are economically and logistically feasible methods that can be used to assess the health of quarry wetlands. Measuring DI is a relatively simple process that could be employed periodically throughout the year at mining sites around the world. Coupled with the water quality data that are often already collected on site, levels of DI could provide an early warning indicator of potential issues, allowing mitigation strategies to be implemented prior to amphibian population declines. As researchers elsewhere adopt DI as a fitness proxy, these data can be used to compare levels observed in quarry wetlands, and ultimately, used to evaluate population health. Overall, the methods outlined here provide a powerful, yet simple, tool to assess the overall health of quarry wetlands that could be easily adopted and implemented at quarries anywhere in the world.

The creation of wetlands during and following mining activities is inevitable. Given the rate at which natural wetlands are being lost and degraded, the ecological importance of these wetlands has never been more important. Here, I show that quarry wetlands within the Kables Sands plant are not only attracting amphibians, but also providing conditions conducive to high individual fitness. Relative to other wetlands in this region, quarry wetlands consistently had greater species richness, as well as individuals with lower levels of developmental instability. Several management practices could further enhance the role quarry wetlands play in promoting amphibian biodiversity, such as planting out emergent zones with native vegetation, and creating a mosaic of wetland types. Implementing these basic survey techniques at quarries is an economically and logistically feasible strategy to ensure quarry wetlands are and remain high-quality systems capable of enhancing local and regional biodiversity and conservation goals.