Abiotic and Biotic Factors Influencing the Assemblage of Tadpoles and Adult Anurans in Coastal Wallum Habitats of Eastern Australia

Author Simpkins, Clay Alan

Published 2013

Thesis Type Thesis (Masters)

School Griffith School of Environment

DOI https://doi.org/10.25904/1912/122

Downloaded from http://hdl.handle.net/10072/367686

Griffith Research Online https://research-repository.griffith.edu.au

Abiotic and biotic factors influencing the assemblage of tadpoles and adult anurans in coastal wallum habitats of eastern Australia

Clay Alan Simpkins, BSc.

Submitted in fulfilment of the requirements for the degree of Master of Philosophy

Environmental Futures Centre, Griffith School of Environment, Griffith University, Gold Coast, Queensland, Australia

Submitted November 2012

Statement of originality

This work has not previously been submitted for a degree or diploma at any university. To the best of my knowledge, the thesis contains no material previously published or written by another person except where due reference and acknowledgement is made in the thesis itself.

Clay Alan Simpkins 16th November 2012

Abstract

The emergence of the global amphibian crisis has seen the extinction of 122 species worldwide, with 18.8% of Australia’s 213 amphibian species being threatened. Despite these declines, little is known about the biology and ecology of certain Australian threatened species. Hence, successful conservation and management of threatened amphibian species cannot be fully realised.

Several environmental variables may influence amphibian adult or tadpole assemblages. These variables include, but are not limited to, water chemistry factors (i.e. pH, salinity, turbidity), predation, competition, hydroperiod and water flow. These variables will influence individual species differently, with each species displaying differences in tolerance to these specific variables.

The coastal wallum vegetation along the eastern coast of Australia is the primary habitat for four specialist frog species (Litoria olongburensis, Litoria freycineti, Litoria cooloolensis and Crinia tinnula) that are listed as Vulnerable under the IUCN Red List. All species are referred to as ‘acid’ frogs due to their association with low pH waters. ‘Acid’ frog populations within protected areas are believed to be stable. However, populations of ‘acid’ frogs occurring outside of protected areas are at risk from ongoing habitat loss and fragmentation. It is therefore vital that conservation managers know which environmental factors influence ‘acid’ frogs to ensure these environmental variables remain constant and populations remain stable. Furthermore, it is imperative to determine if these environmental variables are the same within anthropogenic waterbodies and if ‘acid’ frogs utilise anthropogenic waterbodies. This knowledge would assist in the future prioritisation of waterbodies for conservation. However, the factors influencing ‘acid’ frog species tadpole and adult relative abundance and occupancy within protected and non-protected wallum heathland waterbodies have not been reported.

Therefore, this thesis aims to determine what environmental variables influence these assemblages of ‘acid’ and ‘non-acid’ frog species within and around wallum vegetation of eastern Australia, in both natural and anthropogenic waterbodies. iii

Overall, five tadpole and 14 adult amphibian species were found within surveyed wallum heathland. Several environmental variables influenced the relative abundance and occupancy of L. olongburensis tadpoles and adults. For tadpoles, these variables included pH, water depth and turbidity while variables for adults included pH, water depth, salinity and sedge cover. Environmental variables influencing C. tinnula tadpole occupancy included predatory fish, water depth and turbidity. Several environmental variables influenced adults of competitive species such as L. fallax, indicating that this species is a generalist within the surveyed environment.

Water chemistry variables and the adult amphibian assemblage differed between natural and anthropogenic/compensatory waterbodies. The specialist ‘acid’ frog species had higher relative abundance and reproduced predominantly within natural waterbodies. These patterns are explained by the ideal environmental variables for these species in these natural habitats. The lower relative abundance of generalist ‘non-acid’ frog species in natural waterbodies could be explained by their intolerance to environmental variables, such as low pH. It was therefore possible to differentiate between ‘acid’ frog and ‘non-acid’ frog assemblages in waterbodies using multivariate analyses.

The presence of predatory fish did not influence the relative abundance of L. olongburensis tadpoles or adults. However, the relative abundance of predatory fish was either low or absent in waterbodies where L. olongburensis occurred. Additionally, exotic fish have been proposed as influencing the amphibian assemblage more than other native predatory species. However, predation experiments completed in this study showed that native predators had higher or equal predation rates for tadpoles of L. olongburensis, Limnodynastes peronii and Litoria fallax.

This thesis demonstrates that several environmental variables need to be considered when conservation of ‘acid’ frog species (primarily C. tinnula and L. olongburensis) is undertaken. However, if conservation of all amphibian assemblages within and around wallum heathland areas is the objective, then both anthropogenic and natural waterbodies should be conserved.

Table of Contents

Abstract ……………………...... ……………………………………….…... iii Table of Contents …………...... ……………………………………………. v List of Figures …………...... ……………………………………………...... vii List of Tables …………...... ……………………………………………...... x Acknowledgements …………...... ………………………………..………… xii Chapter 1 - Introduction …………………...………………………………………... 1 1.1. The Importance of Amphibians .…………………..………………..………..... 1 1.2. Assemblages and Communities ……………………………………………….. 2 1.3. Factors Influencing Tadpoles and Adult Amphibian Assemblages……………. 3 1.3.1. The Adult Assemblage ……………………………………………………... 3 1.3.2. Water Quantity and Chemistry …………………………………………….. 5 1.3.2.1. Water Quantity / Hydroperiod …………...... …………………………... 5 1.3.2.2. Water Chemistry ………………………………………………………... 6 1.3.2.2.1. Water pH ……………………………………………………………... 7 1.3.2.2.2. Natural Organic Acids (NOA) ……………………………………….. 8 1.3.2.2.3. Salinity ……………………………………………………………….. 8 1.3.2.2.4. Turbidity and Eutrophication ………………………………………… 9 1.3.2.2.5. Dissolved Oxygen ……………………………………………………. 10 1.3.2.2.6. Water Temperature …………………………………………………… 10 1.3.3. Competition ………………………………………………………………… 10 1.3.4. Predation …………………………………………………………………… 12 1.4. Study Area and Study Species ………………………………………………… 15 1.4.1. Study Area ……………………………………………………………….. 15 1.4.2. Study Fauna ……………………………………………………………… 16 1.4.2.1. Litoria olongburensis …………………………………………………… 19 1.4.2.2. Crinia tinnula …………………………………………………………… 21 1.5. Study Aims ………………………………………………………………….. 24 1.6. References ……………………………………………………………………... 25 Chapter 2 - Environmental variables associated with the distribution and occupancy of tadpoles in naturally acidic, oligotrophic waterbodies……………………….. 39 2.1. Abstract ………………………………………………………………………... 39 2.2. Introduction ……………………………………………………………………. 40 2.3. Methods ………………………………………………………………………... 42 2.4. Results …………………………………………………………………………. 49 2.5. Discussion ……………………………………………………………………... 57 2.6. References …………………………………………………………………..…. 63 Chapter 3 - Battling habitat loss: Suitability of anthropogenic waterbodies for amphibians associated with naturally acidic, oligotrophic environments………. 69 3.1. Abstract ………………………………………………………………………... 69 3.2. Introduction ……………………………………………………………………. 70 3.3. Methods ………………………………………………………………………... 72 3.4. Results …………………………………………………………………………. 77 v

3.5. Discussion ……………………………………………………………………... 85 3.6. References …………………………………………………………………..…. 91 Chapter 4 - Compensatory ponds provide poor habitats for the conservation of frogs associated with naturally oligotrophic, acidic environments…………………… 96 4.1. Abstract …..……………………………………………………..……………... 96 4.2. Introduction ……………………………………………………………………. 97 4.3. Methods ………………………………………………………………………... 98 4.4. Results …………………………………………………………………………. 102 4.5. Discussion ……………………………………………………………………... 108 4.6. References …………………………………………………………………..…. 116 Chapter 5 - Comparison of predation rates between the introduced mosquito fish (Gambusia holbrooki) and native aquatic predators on L. olongburensis, L. fallax and Limnodynastes peronii tadpoles…………………………………… 120 5.1. Abstract ………………………………………………………………………... 120 5.2. Introduction ……………………………………………………………………. 121 5.3. Methods ………………………………………………………………………... 122 5.4. Results …………………………………………………………………………. 126 5.5. Discussion ……………………………………………………………………... 128 5.6. References …………………………………………………………………..…. 135 Chapter 6 - General Conclusions ...... 140 6.1 Chapter Overviews ...... 140 6.1.1 Chapter 2 – Variable influencing wallum heathland tadpole assemblages...... 140 6.1.2 Chapter 3 – Usage of anthropogenic waterbodies and variables influencing adult and amphibian assemblages...... 141 6.1.3 Chapter 4 - Compensatory pond usage by wallum heathland amphibians and variables influencing adult amphibian assemblages ...... 142 6.1.4 Chapter 5 – Predation experiments with G. holbrooki and natural predators ...... 143 6.2. Management Outcomes ...... 143 6.3. Future Priorities for Research ...... 145 6.4. References ...... 146 Chapter 7 - Appendices: Publications on ‘acid’ frogs published during candidature... 144 Appendix 1: Long-range movement in the rare Cooloola sedgefrog Litoria cooloolensis………………………………………………………………...……….. 147

List of Figures

Figure 1.1: The distribution of the ‘acid’ or ‘wallum’ frog species as indicated by red circles. Regional boundaries are indicated by grey lines. State and territory boundaries indicated by solid black circles. Records sourced from the Australian Museum, Queensland Museum, South Australian Museum, Environmental Protection Agency/Queensland Parks and Wildlife Service WildNet database, New South Wales Dept of Environment and Conservation Wildlife Atlas database, and various biologists. Figure obtained from Meyer et al. 2006...... 17

Figure 1.2: The four Australian ‘acid’ frog species – 1. Litoria olongburensis; 2. Crinia tinnula; 3. Litoria freycineti; 4. Litoria cooloolensis……...... 20

Figure 1.3: Distribution of L. olongburensis as indicated by red and blue circles. Red circles indicate records obtained between 1995-2004.Blue circles indicate records obtained before 1995. Regional boundaries are indicated by grey lines. Solid line represents the Queensland / New South Wales state boundary. Records sourced from EPA/QPWS, NSWDEC, the Australian Museum, Queensland Museum, South Australian Museum, and various biologists. Figure obtained from Meyer et al. 2006...... 22

Figure 1.4: Distribution of C. tinnula as indicated by red and blue circles. Red circles indicate records obtained between 1995-2004.Blue circles indicate records obtained before 1995. Regional boundaries are indicated by grey lines. Solid line represents the Queensland / New South Wales state boundary. Records sourced from EPA/QPWS, NSWDEC, the Australian Museum, Queensland Museum, South Australian Museum, and various biologists. Figure obtained from Meyer et al. 2006...... 23

Figure 2.1: Localities of survey sites, with numbers representing the following localities: 1 – Cooloola Section of the Great Sandy National Park; 2- Noosa National Park; 3 – Mooloolah National Park; 4 – Beerwah Scientific Reserve; 5 – Tyagarah Nature Reserve; 6 – Lennox Heads; 7 – Bunjalung National Park; 8 – Yuragir National Park (North); 9 – Yuragir National Park (South). Black dots represent Litoria olongburensis record localities from EPA/QPWS, NSWDEC, the Australian Museum, Queensland Museum, South Australian Museum, and various biologists (Meyer et al. 2006). Solid lines represent Australian coastline and the Queensland / New South Wales state border. Map of Australia shows enlarged area within the rectangle, with solid lines representing the Australian coastline and the Australia’s state and territory borders...... 43

Figure 2.2: ‘Jitter’ plots for quantile regressions of the 0.85(a) and 0.65(b) quantiles (solid line) and the 95% confidence intervals (dotted lines) for mean waterbody pH vii and relative abundance of Litoria olongburensis tadpoles. Circles represent waterbody transects...... 54

Figure 2.3: ‘Jitter’ plots for quantile regressions of the 0.85(a) and 0.65(b) quantiles (solid line) and the 95% confidence intervals (dotted lines) for mean waterbody depth and relative abundance of Litoria olongburensis tadpoles. Circles represent waterbody transects...... 55

Figure 3.1: Amphibian species richness and species presence within each waterbody type. Colours/patterns indicate individual species...... 79

Figure 3.2: Proportion of natural and anthropogenic waterbodies occupied for each recorded anuran species. Records are combined for both visual and acoustic records...... 80

Figure 3.3: nMDS ordination of waterbodies for anuran species where a relative abundance measurement was calculated. Stress associated with 4 dimensions used in MDS ordination was 0.0268. Species ordinations are overlaid. Environmental variables significantly influencing the community structure are displayed. Circles represent waterbodies...... 81

Figure 3.4: ‘Jitter’ plot for relative abundance counts of (a) L. olongburensis and (b) L. fallax in natural and anthropogenic waterbodies. Abbreviations on the x-axis represent the first surveys at natural (NW1), artificial lakes (AL1), road side ditches (RD1) and golf course waterbodies (GCW1) and the second surveys at natural (NW2), artificial lakes (AL2), road side ditches (RD2) and golf course waterbodies (GCW2)...... 83

Figure 4.1: nMDS ordination of amphibian species composition using Axis 1 and 2 from the MDS amphibian species abundance matrix. Black dots represent compensatory ponds while white dots represent established ponds. Species positions within the matrix are displayed ...... 105

Figure 4.2: Gradient analysis using average pH as a gradient with abudance of each species recorded across the survey period. N represents a natural pond while C represents a compensatory pond...... 106

Figure 5.1: Percentage of predators that consumed (black bars) or attacked (white bar) Litoria olongburensis tadpoles for experiments where one individual L. olongburensis was used in each experiment. Number of replicates/experiments is presented above each predatory species...... 124

viii

Figure 5.2: Number of tadpoles consumed for each predatory species. Symbolys represent the number of tadpoles consumed for an individual experiment. ‘o’ represents Limnodynastes peronii, ‘Δ’ represents small Litoria fallax, ‘x’ represents large L. fallax and ‘+’ represent L. olongburensis...... 127

List of Tables

Table 1.1: ‘Acid’ frog conservation status from Queensland, New South Wales and Australian legislation and the IUCN Red List. V = Vulnerable; NT = Near Threatened; - = no status; N/A = not applicable (species not occurring within the state of the Act). (Adapted from Meyer et al. 2006)...... 18

Table 2.2: Comparison of waterbody characteristics associated with the relative abundance and occupancy of L. olongburensis or C. tinnula tadpoles in eastern

Australia. Aikiki models with Δi values less than 4 are presented. + indicates a positive relationship while – indicates a negative relationship to L. olongburensis or C. tinnula tadpole relative abundance or occupancy. Variables with a 2 indicate a unimodal distribution with L. olongburensis or C. tinnula tadpole relative abundance or occupancy...... 51

Table 2.3: Relative importance of waterbody characteristics associated with the relative abundance and occupancy of L. olongburensis or C. tinnula tadpoles in eastern Australia. Model averaged coefficients and relative importance of each environmental predictor for models where Δi < 4 for L. olongburensis relative abundance and occupancy and C. tinnula occupancy are displayed...... 52

Table 2.4: Coefficients of the 0.85 and 0.65 regression quantiles where the independent factors were mean pH and mean water depth. Litoria olongburensis tadpoles were the dependant factor within the regression quantile models...... 56

Table 3.1: Measured variable averages and ranges between the four waterbody types surveyed and for waterbodies with L. olongburensis and L. fallax...... 78

Table 3.2: Correlations (R2 values) between nMDS axis 1 and 2 and environmental variables influencing assemblage structure, with significant correlations (Pr (> r)) highlighted in bold...... 82

Table 3.3: Models with a Δi value < 4 for L. olongburensis and L. fallax adult relative abundance per metre for 2011 surveys. (+) indicates a positive relationship while (-) indicates a negative relationship between relative abundance and the model variable...... 85

Table 3.4: Estimates for model averaged coefficients, standard error (SE), confidence interval (CI) and relative variable importance (RI) for each parameter in models where Δi < 4 for L. olongburensis and L. fallax tadpole relative abundance. (+) indicates a positive relationship while (-) indicates a negative relationship between relative abundance and the model variable...... 86

Table 4.1: Total number of individuals per species detected over the survey period for compensatory and established waterbodies. * indicate threatened species and ^ indicate introduced species listed under the Australian EPBC Act 1999...... 104

Table 4.3: Models with a Δi value < 4 for L. olongburensis and C. tinnula calling activity and relative abundance. + indicates a positive relationship while – indicates a negative relationship to L. olongburensis or C. tinnula calling activity for the variable within the model...... 109

Table 5.1: Number of experiments conducted for each tadpole predator species for multiple prey experiments...... 125

Table 5.2: Average number of tadpoles consumed for each predator species during multiple prey experiments...... 128

Acknowledgements

Numerous people need to be acknowledged for assistance throughout the duration of this thesis. It is difficult to put in words my appreciation for those following people who contributed towards my study. However, I shall attempt to express my appreciation.

I thank my primary supervisor, Associate Professor Jean-Marc Hero, for his advice and sharing his knowledge throughout this entire process. I would especially like to thank my associate supervisor, Dr Guy Castley, for providing valuable guidance and support throughout my entire candidature and for providing invaluable feedback on the final draft of this thesis.

Special thanks go out to my family. To my aunty, Elaine Emery, who provided cheap accommodation in her townhouse and to Mum and Dad for providing a roof over my head when I was without a scholarship. I would also like to give a special thanks to my partner’s parents who provided weekly Sunday dinners and for putting up with my ‘froggish’ antics. Additionally, I thank my partner, Amanda Winzar, who helped me through this process when my morale was low and gave me the incentive to ‘slug it out’ with special ‘slug it out’ brownies, cookies and general ‘bad for you but it tastes so good’ food.

I would like to thank my field assistants – Jodie-Lee Hills, Chays Ogston, Chris Dahl, Diana Virkki, James Bone, Chris Tuohy, Tempe Parnell, Billy Ross, Matt Davies, Donna Treby, Katrin Lowe, Alan Kerr, Gregory Lollback, Nick Clarke and Amanda Winzar. Special thanks go to Jon Shuker for assistance in the field. Without Jons ‘bush bashing’ abilities in the wallum heathland I am unsure if this study would have been possible. Together, Jon and I surveyed the entire distributional range of an amphibian species and, without each other, probably would have succumbed to insanity. I also wish to thank Alan Kerr from the Bribie Island Environmental Protection Society for providing accommodation during fieldtrips.

Michael Arthur, Jon Shuker, Clare Morrison, Gregory Lollback, Donna Treby, Katrin Lowe, Diana Virkki, Ryan Hughes, James Bone, and Sonya Clegg all provided valuable xii advice on either earlier drafts of this thesis, or on statistical issues (primarily on how to use the ‘R’ statistical program (lovingly known as the ‘Pirate Stats Program’)).

I would like to thank Margie Carsburg and Belinda Hachem for their assistance with administrative matters that arose throughout the duration of this thesis. I would also like to thank John Robertson for general guidance and always asking how I was doing. A big thank-you to Jutta Masterton who helped with obtaining field survey equipment – even when the equipment was meant to be forever ‘dead’. I also thank numerous School of Environment staff, including, but not limited to, Tony Carroll, James Furse, Catherine Pickering, Clare Morrison, Sonya Clegg and Hamish McCallum, who were there to listen to my problems, concerns and dilemmas.

I wish to thank the funding bodies that made this thesis possible. Firstly, FKP Pty. Ltd., who provided funding for data collection that contributed towards Chapter 2. Secondly, the Griffith School of Environment, who provided funding for data collection for the remaining thesis chapters. For half of my candidature, I also received a living allowance through the Australian Postgraduate Award scheme.

xiii

1.0 Introduction

The emergence of the global amphibian crisis has seen the disappearance of 122 species of amphibians (Stuart et al., 2004), with 18.8% of Australia’s 213 species being threatened (Hero and Morrison, 2004). Despite these declines, little is known about the population dynamics, biology and ecology of certain Australian threatened species (Hines et al., 1999; Hero et al. 2006).

Understanding what environmental variables influence amphibians within the landscape is essential if conservation management is to be conducted successfully. Since most amphibians occur in different ecological niches during different stages of their lifecycle (Wells, 2007) it is imperative to determine what environmental factors influence amphibian distributions during all lifecycle stages.

1.1 The Importance of Amphibians

Amphibian reproductive modes are numerous, with amphibian larvae developing in both the aquatic and terrestrial environment (Haddad and Prado, 2005; Wells, 2007). The non-reproductive larval stages (tadpoles) occur in different ecological niches compared with the adult stages (Wells, 2007; Halliday, 2008; McDiarmid and Altig, 2010). The tadpole stage is pivotal within the amphibian lifecycle and has been described as having ‘the potential to have the greatest impact on the continuing persistence of the (amphibian) population’ (Lane and Mahony, 2002).

Tadpole composition and abundance heavily influence the structure of many aquatic communities. Sediment dynamics (Flecker et al., 1999; Ranvestal et al., 2004) and the assemblage and abundance of algae (Morin, 1995; Ranvestal et al., 2004) and zooplankton (Mokany, 2007) are influenced by the assemblage and abundance of tadpoles. Therefore, freshwater aquatic communities rely heavily upon amphibian tadpoles in maintaining ecosystem equilibrium.

Adult anurans also play a key role in ecosystem function as they are food sources for numerous predators and prey on numerous fauna (Duellmann and Trueb, 1986; Wells, 2007; Crump, 2010). The adult assemblage is also important in determining the tadpole assemblage as tadpoles cannot occur in areas where adults fail to deposit eggs. Additionally, human society has benefited from

1 amphibians with the discovery and isolation of chemical compounds from adult anurans (Crump, 2010). Therefore, the importance of amphibians to human society and environmental processes cannot be underestimated.

Amphibian adult density has been linked to larval survivorship in certain species. For example, densities of adult Rana sylvatica have been found to be dependent on the survival of R. sylvatica larvae (Berven, 1990). This is also evident in Bufo calamita where the density of adults was positively correlated with B. calamita metamorph density (Beebee et al., 1996). Therefore, factors that influence amphibian larval assemblages will also influence the adult amphibian assemblages.

1.2 Assemblages and Communities

For the purpose of this thesis, an assemblage can be described as a group of species that occur within a particular environment where interactions amongst individuals do not have to occur (Retallick, 2000). A community is defined as individuals of different species occurring within a particular environment that interact with each other (Hickman Jr. et al., 1998).

Interactions occurring between individuals within communities may be positive, negative or neutral and can potentially influence a species’ distribution and population structure. For example, fish predation on tadpoles has a negative interaction on the tadpole but a positive interaction for the fish by providing nutrition. These interactions may exclude or reduce specific amphibian species from waterbodies (Kats et al., 1988; Hecnar and M'Closkey, 1997; Hero et al., 1998; Kats and Ferrer, 2003; Vonesh et al., 2009) and thus structure the overall amphibian tadpole assemblage occurring within a waterbody.

Interspecific competition and predation are two important biotic interactions that can structure an assemblage or community (Schoener, 1983), but these are also influenced by other environmental factors. The influence of environmental factors on individual species will differ as species have variable responses to these factors (Cushman, 2006). The environmental effects may also differ between populations of the same species at different spatial scales (Pierce, 1985; Grand and Cushman, 2003). Furthermore, low levels of disturbance are believed to aid in maintaining high species diversity within an assemblage (Death and Winterborn, 1995).

The importance of environmental factors in influencing amphibian assemblage and community structure has been noted within the literature (see 1.3 of this thesis). Arguments that favour multiple factors structuring communities and assemblages are likely correct as assemblages and communities contain multiple species that will co-exist with different predators and competitors and be tolerant to differing ranges of environmental variables.

1.3 Factors influencing Tadpole and Adult Amphibian Assemblages

There are numerous environmental factors that have the potential to influence amphibian adult and tadpole assemblages. These factors include, but are not limited to, predation (Kats et al., 1988; Hecnar and M'Closkey, 1997; Hero et al., 1998; Gillespie and Hero, 1999; Vonesh et al., 2009), competition (Wiltshire and Bull, 1977; Hickman Jr. et al., 1998; Twomey et al., 2008), water chemistry (Gosner and Black, 1957; Pierce, 1985; Freda, 1986; Freda and Taylor, 1992; Smith et al., 2007; Sparling, 2010), water quantity (i.e. hydroperiod) (Wilbur, 1987; Snodgress et al., 2000; Baber et al., 2004; Moreira et al., 2010) and water flow (Richards, 2002). Additionally, factors that influence the adult assemblage may influence the tadpole assemblage and vice versa. Amphibian adult density has been linked to larval survivorship in certain species. For example, densities of adult Rana sylvatica have been found to be dependent on the survival of R. sylvatica larvae (Berven, 1990). This is also evident in Bufo calamita where the density of adults was positively correlated with B. calamita metamorph density (Beebee et al., 1996). These factors may exclude certain species from particular waterbodies or be tolerated at different levels, with the tolerance level depending on the individual species (Gosner and Black, 1957; Pierce, 1985; Freda, 1986; Freda and Taylor, 1992; Meyer, 2004; Smith et al., 2007; Sparling, 2010).

1.3.1 The Adult Assemblage

Vegetation cover (Gibbs, 1998; Girish and Krishna-Murthy, 2009) and habitat size (Kolozsvary and Swihart, 1999) influence the distribution of adult amphibians. Factors influencing adult distributions may also be correlated with other variables ultimately influencing adult distributions. For example, Girish and Krishna-Murthy (2009) found increased light intensity as a result of decreased forest cover affected air and water temperatures. Furthermore, the abundance or 3 emergence of particular vegetation species (Lemckert et al., 2006; Shuker, 2012), or the proportion of the water margin with emergent vegetation (Hazell et al., 2004) Lemckert et al., 2006, may also influence amphibian usage or species abundances within waterbodies. Pond isolation may also negatively influence adult amphibian species richness (Smallbone et al., 2011), individual species usage (reviewed in Marsh and Trenham 2001) or breeding success (Marsh et al., 1999; reviewed in Marsh and Trenham 2001).

Temperature and rainfall will affect the adult contribution towards the tadpole assemblage as calling and breeding may not occur when temperatures and water levels are inadequate (Duellmann and Trueb, 1986; Oseen and Wassersug, 2002; Wells 2007). For example, at high altitudes or in temperate environments, adult amphibians will often hibernate when temperatures fall below ideal conditions (Pearson and Bradford 1976; Carey 1978; Pinder et al., 1992; Wells 2007), while calling and the selection of spawning sites will peak when optimal water temperatures and depth are reached (Oseen and Wassersug, 2002; Goldberg et al., 2006).

Adult females have the ability to influence the tadpole assemblage by choosing oviposition sites and the number of eggs that are deposited (Resetarits Jr and Wilbur, 1989). Furthermore, by avoiding waterbodies that contain predatory cues when depositing eggs (Binckley and Resetarits Jr., 2003; Orizaola and BraÑa, 2003), females may influence the tadpole assemblage. Site selection can also be influenced by the female’s ability to detect competitors (Resetarits Jr and Wilbur, 1989) and intolerant water chemistry levels (Haramura, 2008).

It is imperative to note that, despite the importance of the adult assemblage, the presence of adults does not indicate a site of reproduction (Mazerolle, 2005) and thus tadpoles of the adult species recorded at a waterbody may be absent. Additionally, Girish and Krishna-Murthy (2009) found factors influencing the occurrence of tadpoles also influenced adult occurrence. Therefore, scenarios and environmental factors that affect the tadpole assemblage may also influence the adult assemblage.

1.3.2 Water Quantity and Chemistry

Amphibians are largely dependent on water throughout all stages of their lifecycle (Halliday, 2008; Sparling, 2010), with the egg and larval stages of numerous amphibian species spent within the aquatic environment (Duellmann and Trueb, 1986; McDiarmid and Altig, 2010). Adult stages are not restricted to waterbodies and have the freedom to move within the terrestrial landscape (Johnson et al., 2007; Simpkins et al., 2011). Nevertheless, water is an essential commodity in both quantity (hydroperiod) and quality (chemistry).

1.3.2.1 Water Quantity / Hydroperiod

The time water is present in a waterbody (hydroperiod) influences the structure of anuran tadpole assemblages either directly (Wilbur, 1987; Snodgress et al., 2000; Baber et al., 2004; Moreira et al., 2010), or indirectly (Herrmann et al., 2005). Hydroperiod may indirectly influence amphibian larval assemblages by influencing predator and competitor assemblages and abundance (Woodward, 1983; Richter-Boix et al., 2007), or water chemistry variables (Herrmann et al., 2005).

Hydroperiod influences assemblage and abundance via natural selection for species that can successfully reproduce in temporary, permanent or both types of waterbodies. For example, ephemeral amphibian breeders are often excluded from permanent waterbodies due to the presence of different predator assemblages (Welborn et al., 1996; Hero et al., 1998; Richter-Boix et al., 2007) and, potentially, predator size, which is often larger in permanent waterbodies (reviewed in Welborn et al., 1996; Richter-Boix et al., 2007). Ephemeral breeders, therefore, need to metamorphose quickly before aquatic predators can become established. Overall success is therefore higher in species that arrive in ephemeral waterbodies early (Wilbur, 1997). The opposite may occur for permanent breeders when they are excluded from ephemeral waterbodies due to shorter hydroperiod lengths and inadequate metamorphosis time (Skelly, 1995). The effects of aquatic predators on structuring tadpole assemblages are outlined in the ‘Predation’ section of this chapter.

In addition to hydroperiod, a decrease in water level can result in a decrease in food availability (Loman, 1999). This may be problematic in ephemeral waterbodies where limited food resources could restrict development (Newman 1994), as tadpoles need to metamorphose before waterbody desiccation occurs. Tadpoles occurring in ephemeral waterbodies have evolved to metamorphose early in response to decreasing water levels (Lane and Mahony, 2002; Márquez-García et al., 2010), despite a reduction in resources (Loman, 1999). Additionally, temperature influences the developmental rate of tadpoles (Duellmann and Trueb, 1986; Orizaola and Laurila, 2009), with temperature potentially being elevated in ephemeral waterbodies (Noland and Ultsch, 1981). Early metamorphosis ensures the short-term survival of the individual but often results in decreased juvenile body size, potentially affecting the individual’s long-term survival (Smith, 1987; Lane and Mahony, 2002; Altwegg and Reyer, 2003). To avoid early metamorphosis and waterbody desiccation anurans that breed within ephemeral waterbodies often breed after extensive rainfall when water levels are sufficient to ensure optimal chances of metamorphosis.

Shorter hydroperiods and decreasing water levels also increase the amount of UV-B radiation. Exposure to UV-B radiation increases fungal infection in eggs where amphibians were forced to deposit eggs in waterbodies with lower than average water levels (Kiesecker et al., 2001). Finally, water chemistry also changes in response to varying hydroperiod lengths as reported by Herrmann et al. (2005), where conductivity was significantly lower in ephemeral waterbodies when compared with permanent waterbodies. This is just one example of the effects of water chemistry that are outlined in further detail below.

1.3.2.2 Water Chemistry

The permeability of anuran skin facilitates the uptake of water through osmosis (Shoemaker and Nagy, 1977; Bentley and Yorio, 1979), which in turn is influenced by differing levels of particular water chemistry variables (Wells, 2007). Water chemistry is therefore important for amphibian larval and embryo survival.

Water chemistry factors can be tolerated at different levels (tolerance range) depending on the anuran species (Gosner and Black, 1957; Pierce, 1985; Freda and Taylor, 1992; Chinathamby et al., 2006; Persson et al., 2007; Smith et al., 2007; Rios-López, 2008; Barth and Wilson, 2010;

Sparling, 2010) and the lifestage of the individual (Strahan, 1957; Freda, 1986). The tolerance range can be categorised into sub-lethal, lethal or non-lethal/optimal with the effects varying between and within species. Water chemistry can also influence anuran tadpole assemblages by influencing parasite, predator and competitor assemblages (Sparling, 2010).

Water chemistry factors that may include conductivity (caused by the quantity of anions and cations) (Smith et al., 2007; Sparling, 2010), pH (caused by hydrogen ion concentration) (Sparling, 2010), salinity (concentration of chloride salts) (Sparling, 2010), turbidity, (influenced by particle suspension of inorganic and organic matter) (Sparling, 2010), dissolved oxygen (Sparling, 2010) and natural organic acids (Steinberg et al., 2006). Dissolved pollutants have also been shown to affect anuran tadpole assemblages (Sparling, 2010). These factors may influence amphibian assemblages and are further discussed in the sections below.

1.3.2.2.1 Water pH

Amphibians that breed successfully in acidic aquatic environments can tolerate lower pH levels than those breeding in non-acidic environments (Gosner and Black, 1957; Freda, 1986; Meyer, 2004). Despite this, amphibians have been known to breed in water bodies that were either outside or near their pH tolerance level where rainfall temporarily increases pH levels (Sadinski and Dunson, 1992). Additionally, intraspecific tolerance may differ amongst populations (Pierce, 1985; Glos et al., 2003; Persson et al., 2007). For example, populations of Rana sylvatica, an anuran tolerant of low pH levels, differ in their pH tolerance range depending on the level of acidity exposure, which varies across geographical locations (Pierce, 1985).

Sub-lethal pH levels can produce tadpoles hatching with abnormalities (Gosner and Black, 1957; Andrén et al., 1988), increase time to metamorphosis, reduction in body size (Cummings, 1986) and, indirectly, a reduction in clutch and egg size (Räsänen et al., 2008). Lethal effects of pH can result in tadpoles failing to hatch from eggs (Gosner and Black, 1957; Sadinski and Dunson, 1992; Meyer, 2004), or inhibit fertilisation due to reduced movement or death of sperm (Schlichter, 1981). However, jelly membranes on eggs may also act as a buffer to acidic waters (Picker et al., 1993). To combat these effects some amphibian tadpoles have adapted mechanisms to detect pH levels and can actively avoid exposure to unsuitable pH levels (Freda and Taylor, 1992). Certain

7 low pH waterbodies can be heterogeneous, in relation to pH (Freda and Taylor, 1992), and thus detection of unsuitable pH levels may influence the distribution of tadpole species within waterbodies. Additionally, tolerance to varying pH levels increases with increased development stages (Pierce, 1985; Freda, 1986).

1.3.2.2.2 Natural Organic Acids (NOA)

Naturally occurring organic acids (NOA) are derived from decomposing organic matter and are often referred to as dissolved humic substances. The dark brown coloration that occurs in waterbodies in wallum heathland of Eastern Australia and in the ‘blackwaters’ of Rio Negro in the Amazon are attributed to the chemical properties of the waters; that is low or absent in magnesium and/or calcium (soft-water), low buffering capacity and high organic acids (Barth and Wilson, 2010). In some isolated waterbodies, natural organic acids can decrease pH levels (Sparling, 2010).

It is also believed that NOA can influence the faunal community independent of pH (Steinberg et al., 2006), however, current knowledge on how humic/organic acids influence tadpole communities is lacking (Barth and Wilson, 2010). While hatching success can be reduced in low pH waters with high levels of NOA, this may be dependent on individual species tolerance levels (Picker et al., 1993). Within soft-water, humic substances can either protect (Wood et al., 2003; Steinberg et al., 2006), or expose (Steinberg et al., 2006), other non-amphibian aquatic fauna (i.e. fish) to ion-loss. Under these conditions humic substances can impose negative stresses on aquatic invertebrates (Timofeyev et al., 2006). Therefore, NOA have the potential to structure tadpole assemblages directly, by selecting for tadpole species that can tolerate high levels of NOA within the waterbody, or indirectly, by influencing the structure of the aquatic predator assemblage.

1.3.2.2.3 Salinity

Amphibians are rarely detected in waters with high salt concentrations due to their inability to efficiently osmoregulate under these conditions (Gomez-Mestre et al., 2004). Despite the majority of amphibians being intolerant to saline waters some can live in waterbodies with salinity levels close to seawater (e.g. Fejervarya cancrivora (crab-eating frog)). Adult F. cancrivora regulate the

8 osmotic process in response to salinity by increasing the amount of urea, sodium and chloride present within their bodies (Wells, 2007). Tadpoles of this species displayed signs of increased sodium and chloride but regulated the osmotic process by excreting salts via the gill membranes (Wells, 2007). A similar increase in sodium, chloride and calcium present within tadpoles in saline waters was observed in Bufo calamita, although there was no mention of salt being excreted across the gill membranes (Gomez-Mestre et al., 2004).

Salinity tolerance levels will differ among species (Smith et al., 2007). Sub-lethal levels of water salinity can cause an increased time to metamorphosis (Christy and Dickman, 2002; Gomez- Mestre et al., 2004; Chinathamby et al., 2006), a reduction in body weight (Christy and Dickman, 2002; Gomez-Mestre et al., 2004) and retardation of external features (Rios-López, 2008). Lethal effects of salinity can cause death to individual tadpoles and failure to metamorphose (Christy and Dickman, 2002; Chinathamby et al., 2006). The lethal effects of salinity can decrease with older developmental stages in some tadpoles (Strahan, 1957). To overcome the effects of sub-lethal and lethal levels of salinity coastal frog species (e.g. Buergeria japonica) have evolved the ability to detect water salinity levels and will actively choose their oviposition sites in non-saline water (Haramura, 2008).

1.3.2.2.4 Turbidity and Eutrophication

The effects of turbidity on tadpoles are not well known (Schmutzer et al., 2008), but may influence tadpole assemblages by decreasing predation levels in turbid water. For example, tadpoles of Phrynomantis microps increased their schooling density and size when waters were less turbid and this was attributed to an increased risk of predation in clearer water (Spieler, 2003).

Other studies have shown that nitrate levels can influence the survival of amphibian larvae by inhibiting growth and development (Mann and Bidwell, 1999). Therefore, Meyer et al. (2006) proposed that increasing nitrate levels in waterbodies along Australia’s eastern seaboard (i.e. nutrient poor wallum heathland waterbodies) could alter the viability of that habitat for ‘native’ species. Additionally, the increase in nitrates could alter vegetation community structure towards a vegetation community that would be unsuitable for native species (Meyer et al., 2006).

1.3.2.2.5 Dissolved Oxygen

Dissolved oxygen (D.O.) levels differ depending on the waterbody in question. Well mixed lotic waterbodies are usually higher in D.O. compared to lentic waterbodies where D.O. usually decreases within increasing depth (Sparling, 2010). Therefore, tadpoles living in low oxygenated (hypoxic) waterbodies rely on oxygen being taken up primarily through their lungs by surfacing to the top of the water (Wassersug and Seibert, 1975; Sparling, 2010). Surfacing can come at a cost as it increases predation risk from visually orientated predators (Feder, 1983).

The proportion of dissolved oxygen needed for tadpole survival differs among species (Sparling, 2010). Furthermore, while some anuran eggs hatch earlier when exposed to low oxygenated waters, low oxygen can also be lethal and result in death of eggs (Seymour et al., 2000).

1.3.2.2.6 Water Temperature

Amphibians lack the ability to produce their own body heat. Therefore, environmental temperature is extremely important for behaviour, metabolic rates and other physiological processes (Sparling, 2010). As mentioned previously, water temperature can influence the developmental process and growth rates of tadpoles (Duellmann and Trueb, 1986; Orizaola and Laurila, 2009). Low temperatures will often result in slow development (Duellman and Trueb, 1986) and therefore be a key factor for tadpole survival in ephemeral waterbodies where hydroperiod can be short. Additionally, water temperature may influence oviposition timing and location (Goldberg et al., 2006). Water temperature may also influence levels of D.O. with concentrations decreasing as water temperatures increase (Sparling, 2010).

1.3.3 Competition

Darwin considered competition an important factor in ecological communities as it reduces fitness of the weaker competitor (Hickman Jr. et al., 1998). Competition can occur within (intraspecific) or amongst (interspecific) species when a resource being shared is scarce and will result in

10 competitive exclusion or competitive coexistence (Hickman Jr. et al., 1998; Twomey et al., 2008), thereby structuring communities (Wiltshire and Bull, 1977).

Interspecific competition among amphibian species has the potential to alter rates of tadpole growth and development (Morin, 1986; Wilbur, 1987; Relyea, 2004; Twomey et al., 2008), as well as aquatic, non-amphibian species (Mokany and Shine, 2003). Slower growth rates can be lethal in ephemeral waterbodies if metamorphosis does not occur before waterbody desiccation. Furtermore, delaying metamorphosis may also prolong exposure to aquatic predators and increase the predator-prey interaction time.

High competition can result in smaller body size at metamorphosis (Semlitsch and Ryer, 1992; Rudolf and Rödel, 2007) and can lead to higher mortality as a metamorph, lower reproductive success as an adult (Lane and Mahony, 2002; Altwegg and Reyer, 2003) or increased time to sexual maturity (Smith, 1987; Scott, 1994). The reduction in growth and development under competition can potentially be related to food availability which would be lower under increased competition. Decreased food availability has also been shown to negatively influence the growth of tadpoles (Griffiths et al., 1993; Mokany and Shine, 2003) and the size at metamorphosis (Newman, 1994).

Interspecific competition can be minimised both spatially and temporally within the tadpole assemblage. Temporally, the phenology of anurans contributes towards reducing interspecific competition when eggs are deposited at different time intervals (Heyer, 1973; Toft, 1985; Wells, 2007; Crossland et al., 2009). Larger tadpoles have been noted to outcompete and kill smaller tadpoles (Toft, 1985) and earlier hatching can provide older tadpoles with first choice of resources and the potential to outgrow their competitors. Additionally, hatching at different times will minimise interaction with other species and reduce resource competition.

The spatial structuring of the tadpole assemblage in different columns of a waterbody can remove or minimise interspecific competition (Heyer, 1973). Predation can additionally reduce inter- and intraspecific competition by minimising the number of individuals present (Wilbur, 1997) and removal of species that are susceptible to predation (Kats et al., 1988). Female choice of

11 oviposition sites will further reduce the risk of competition with some species actively avoiding sites where more competitive species are present (Resetarits Jr and Wilbur, 1989).

Morphology of tadpoles may change under interspecific competition. To maximise success of obtaining food, tadpoles of Rana sylvatica and Rana pipiens, have increased their mouth width by up to 10% and 5%, respectively, when competing with each other (Relyea, 2000). Alternatively, tadpoles may shift their dietary preference to reduce competition of food resources through phenotypic plasticity (Pfennig and Murphy, 2002). Therefore, tadpoles which display lower levels of phenotypic plasticity are less likely to adapt to competitive pressures, having an increased chance of exclusion.

Ultimately, tadpoles should ideally occur in waterbodies that maximise their ability to grow, develop and avoid any negative interactions that results from competition (Retallick, 2000).

1.3.5 Predation

Predation is often an underlying factor determining tadpole assemblages within waterbodies (Kats et al., 1988; Hecnar and M'Closkey, 1997; Hero et al., 1998; Vonesh et al., 2009) and is arguably the most important biotic factor influencing tadpole assemblages both spatially and temporally (Heyer et al., 1975). Predators of tadpoles include fish (Hero et al., 1998; Baber and Babbitt, 2003; Vonesh et al., 2009), aquatic invertebrates (Heyer et al., 1975; Fox, 1978; Stoneham, 2001; Jara, 2008; Álvarez and Nicieza, 2009), salamanders (Morin, 1995) and other tadpoles (Heyer et al., 1975; Álvarez and Nicieza, 2009).

Amphibian tadpoles have evolved anti-predator strategies to coexist with their natural predators. These include unpalatability or chemical defences (Kats et al., 1988; Hero et al., 2001; Gunzburger and Travis, 2005), behavioural avoidance (Skelly, 1994; Relyea, 2003; Gregoire and Gunzburger, 2008; Saidapur et al., 2009; Smith and Awan, 2009), morphological adaptations (Hecnar and M'Closkey, 1997; McCollum and Leimberger, 1997; Touchon and Warkentin, 2008) and grouping/schooling (Watt et al., 1997). Tadpoles can also detect predators through chemical cues. In a classical co-evolutionary arms race, predators respond to tadpoles’ evolutionary tactics by

12 evolving abilities of their own to overcome the tadpoles’ anti-predator strategies (Hickman Jr. et al., 1998; Brodie III and Brodie Jr., 1999).

Predatory fish can exclude some species of tadpoles that lack adequate anti-predator strategies from waterbodies (Kats et al., 1988). The level of predation on a tadpole will be determined by a number of factors, including the tadpole’s vulnerability to the predator assemblage and predator abundance (Hossie and Murray, 2010). Predator assemblage and abundance will often vary depending on biotic and abiotic factors (Woodward, 1983; Babbitt et al., 2003). For example, permanent waterbodies will contain predators that require permanent water to survive. Conversely, in ephemeral waterbodies predators may be absent in the ‘starting’ period of the waterbody, or be of a smaller size when compared with predators in permanent waterbodies (Richter-Boix et al., 2007). Thus, it may be beneficial for tadpoles in ephemeral waterbodies to develop quickly due to the risk of predation increasing with time (Duellmann and Trueb, 1986; Relyea, 2007).

If rapid development is not possible tadpoles can co-exist with their predators by using refuges (Babbitt and Tanner, 1998; Kopp et al., 2006; Saidapur et al., 2009). An increase or decrease in use of refuges is often dependant on the type of predator being avoided (Morin, 1986; Smith et al., 2008; Smith and Awan, 2009). For example, the effectiveness of habitat refugia will often depend on the size of the predator, with larger predators being avoided more successfully than smaller predators (Babbitt and Tanner, 1998). Furthermore, predators, like Dytiscus sp (diving water beetles), may adopt different hunting strategies under different levels of habitat complexity and thus use of refuges may not decrease the overall risk of predation (Michel and Adams, 2009). Additionally, refugia are not always beneficial and complex environments have been found to increase predation in fast swimming tadpoles by hindering the tadpoles swimming ability (Saidapur et al., 2009).

Tadpoles may reduce time spent foraging and/or moving to reduce the risk of predation (Richter- Boix et al. 2007, Relyea 2007, Saidapur et al. 2009, Smith and Awen 2009), with altered behaviour often influenced by the threat level that is associated with a predator. For example, experiments performed on Rana sylvatica showed a decrease in movement when exposed to ‘fresh’ predatory cues, but increased when predation threat levels were lower (Ferrari and Chivers 2008). Furthermore, some tadpoles have the ability to learn, which may influence movement

13 behaviour based on past predator experiences (Shah et al. 2010). A decrease in movement/foraging time can potentially come at a cost of metamorphic weight, size and growth rate which can have detrimental effects during later lifecycle stages (Skelly 1995).

Tadpoles may form groups or schools in an attempt to reduce predation (Rödel and Linsenmair, 1997; Watt et al., 1997; Spieler, 2003; Stav et al., 2007). An increase in schooling size will reduce the probability of predation on individuals, despite an increase in the number of overall attacks by predators as schooling size increases (Watt et al., 1997).

The risk of predation is lower for larger tadpoles than small tadpoles (Brodie Jr. and Formanowicz Jr., 1987; Jara, 2008; Arendt, 2009), but is also affected by predator size. Some species also show adaptive plasticity allowing them to grow larger in deeper waterbodies compared to shallow waterbodies (Loman and Claesson, 2003), potentially reducing predation in deeper, more permanent, waterbodies where larger predators occur (Richter-Boix et al., 2007). In the absence of a large body size some tadpoles have evolved the ability to ‘sprint’ (i.e. swim fast), to actively evade predators (Arendt, 2009; Saidapur et al., 2009).

Colouration can also be used as a camouflage defence against predators that see primarily in particular light spectrums (Touchon and Warkentin, 2008), or for camouflage within the natural environment (McCollum and Leimberger, 1997). It has been noted that some tadpole species have the ability to change their colouration when exposed to predators (Touchon and Warkentin, 2008) and can centre colouration to particular areas in an attempt to focus attacks away from important vulnerable areas (Van Buskirk et al., 2004).

Particular tadpole species have been noted to be unpalatable (Heyer et al., 1975; Henrikson, 1990; Lawler and Hero, 1997; Jara and Perotti, 2008) due to secretions that are toxic to particular predators. The level of palatability may also change as tadpoles develop (Lawler and Hero, 1997), however, unpalatable tadpoles may still be attacked and suffer injuries from predators (Jara and Perotti, 2008), that may result in cannibalism from other tadpoles (Álvarez and Nicieza, 2009). Additionally, in an attempt to reduce tadpole movement, tadpole tail-nipping by predatory fish has been recorded (Baber and Babbitt, 2003) in an attempt to allow for easier consumption regardless of tadpole size.

Predators may use visual (Rödel and Linsenmair, 1997; Saidapur et al., 2009), or chemical (Saidapur et al., 2009), cues to hunt prey and thus the effectiveness of one anti-predator defence may not work on all predators within a system (Hero et al., 2001; Saidapur et al., 2009). Decreased palatability, for example, has been known to work against Pyrrhulina sp. for tadpoles of Hyla boans but is ineffective against the odonate naiad Gynacantha membranalis (Hero et al., 2001). To overcome these scenarios a combination of anti-predator defences may be employed by tadpoles (Kats et al., 1988). Different strategies may not occur together but could shift with ontogeny (e.g. a reduction in activity as unpalatbility defence decreases with increase development) (Jara and Perotti, 2008).

Introduction of exotic predators into an aquatic ecosystem can have significant impacts upon the naturally occurring tadpole assemblage (Gillespie and Hero, 1999; Kats and Ferrer, 2003), as tadpoles may not have evolved any effective anti-predator strategies against introduced predators (Kats and Ferrer, 2003). Alternatively, native tadpole species may fail to detect predatory cues from non-native predatory species (Polo-Cavia et al., 2010) and may therefore be heavily predated upon.

Predators may therefore influence the amphibian community by removing amphibian species that lack adequate anti-predator defences and selecting for amphibian species that can co-exist with their predators. Species that lack these defences will be presented with higher exclusion pressures within an ecosystem. Alternatively, temporal separation will select for tadpoles that avoid predators completely.

1.4 Study Area and Study Species

1.4.1 Study Area

The study area encompassed waterbodies within wallum vegetation along the eastern coast of Australia, between Rainbow Beach, QLD, and Wooli, N.S.W. (Figure 1.1). Wallum vegetation occurs along the coastal lowlands of eastern Australia between Newcastle, N.S.W., and Rockhampton, QLD (Griffith et al., 2008). For the purposes of this study ‘wallum’ is described as

15 the vegetation communities that include Banksia woodland, sedgeland, heathland and Melaleuca swamps (Hines et al., 1999; Griffith et al., 2003) occurring within the coastal lowlands of South East Queensland (SEQLD) and north east New South Wales (Griffith et al., 2003). Within the study area wallum vegetation communities occur along ‘dunefields, beach ridge plains and sandy barrier flats’ with soils that are often acidic (pH 3.4-5.4) (Griffith et al., 2008), sandy (Hines et al., 1999) and low in nitrogen and phosphorus (Groves, 1981).

Waterbodies associated with wallum communities are also low in nutrients and contain acidic water (pH <5.5) (Meyer et al., 2006), as a result of decaying detritus/organic matter (Barth and Wilson, 2010). The dark brown colouration that occurs in waterbodies of wallum heathland are attributed to the chemical properties of the waters; low or absent in magnesium and/or calcium (soft-water), low buffering capacity and high organic acids (Barth and Wilson, 2010).

Wallum vegetation communities are highly flammable and fire is believed to play an integral part in certain wallum ecosystems (Specht, 1981). Drought and flooding are also common within wallum communities as seasonal rainfall can fluctuate considerably (Griffith et. al., 2004). Within SE QLD rainfall can be largely dependent on cyclones and thunderstorms (Coaldrake, 1961).

1.4.2 Study Fauna

Occurring within coastal wallum vegetation along the eastern coast of Australia (Figure 1.1) are four species of anurans that have been described as ‘acid’ or ‘wallum’ frogs (Ingram and Corben, 1975; Hines et al., 1999; Meyer et al., 2006). The terming ‘acid’ frog is based around these species ability to occur in waters with low pH (Ingram and Corben, 1975) and include Litoria cooloolensis, Litoria freycineti and, the primary study species, Litoria olongburensis and Crinia tinnula (Meyer et al., 2006) (Figure 1.2).

The main threat facing ‘acid’ frogs is habitat loss for agricultural, residential or infrastructure development (Hines et al., 1999; Meyer et al., 2006). Threat severity is increased as the majority of the ‘acid’ frog species distributions overlap with areas where human population growth rates are highest (Hines et al., 1999). Other threatening processes to ‘acid’ frog populations include habitat

Figure 1.1: The distribution of the ‘acid’ or ‘wallum’ frog species as indicated by red circles. Regional boundaries are indicated by grey lines. State and Territory boundaries indicated by solid black circles. Records sourced from the Australian Museum, Queensland Museum, South Australian Museum, Environmental Protection Agency/Queensland Parks and Wildlife Service WildNet database, New South Wales Dept of Environment and Conservation Wildlife Atlas database, and various biologists. Figure obtained from Meyer et al., 2006.

17 degradation, invasive flora species, disease (i.e. Chytrid fungus), inappropriate fire management, hydrological alteration, and introduction or translocation of fish species (Gillespie and Hero, 1999; Hines et al., 1999; Meyer et al., 2006). Chemicals used to control mosquitoes and weeds may also pose a risk (Meyer et al., 2006). Despite these threats, it is believed that populations of ‘acid’ frogs occurring within protected areas (i.e. National Parks) are stable (Hines et al., 1999; Lewis and Goldingay, 2005).

All ‘acid’ frog species are listed as ‘Vulnerable’ or ‘Near Threatened’ under the Queensland Nature Conservation Act 1992 (NCA 1992) and the New South Wales Threatened Species Conservation Act 1995 (TSC Act 1995). Litoria olongburensis is the only species currently listed under the Commonwealth Environment Protection and Biodiversity Conservation Act 1999 (EPBC Act 1999). Internationally, the World Conservation Union lists the ‘acid’ frog species as either ‘Endangered’ or ‘Vulnerable’ (Table 1.1).

Table 1.1: ‘Acid’ frog conservation status from Queensland, New South Wales and Australian legislation and the IUCN Red List of Threatened Species. V = Vulnerable; NT = Near Threatened; - = no status; N/A = not applicable (species not occurring within the state of the Act). (Adapted from Meyer et al., 2006).

Species NCA1 1992 TSC2 Act EPBC3 Act IUCN4 1995 1999 Litoria olongburensis V V V V

Litoria coolooensis NT N/A - E Litoria freycineti V - - V Crinia tinnula V V - V

1.4.2.1 Litoria olongburensis

One target study species, L. olongburensis, belongs to the Hylidae family and is often referred to as the Wallum Sedge Frog, Olongburra Frog or the Sharp-Snouted Reed Frog (Tyler, 1997; Meyer et al., 2006). The species can range between 25-31mm SVL and varies in coloration, with the dorsum of different individuals being recorded as grey-brown, beige or bright green (Meyer et al., 2006; Lowe and Hero, 2012). Ventral coloration is usually white or cream with the thighs being blue or purple blue with red or orange (Figure 1.2) (Meyer et al., 2006).

On the mainland, distribution of L. olongburensis occurs between Rainbow Beach, QLD, and Woolgoolga, N.S.W. (Figure 1.3) in well vegetated, low nutrient, acidic, tannin stained, ephemeral swamps that consist of reeds, sedges and emergent ferns (Tyler, 1997; Robinson 2002, Meyer et al., 2006) Fragmentation of habitat occurs throughout this species range (Tyler, 1997, Meyer et al. 2006).

Breeding occurs during spring, summer and autumn periods when males call and eggs are deposited at the base of sedges or reed stems (Anstis, 2002; Meyer et al., 2006). Tadpoles can grow to 37mm in tail length and 13mm in body length at Gosner stage 37 (Anstis, 2002). Tadpoles are located in the mid or surface water and are well camouflaged against the tannin stained waters (Anstis, 2002), or can be found foraging or resting on matted sedges (Meyer et al., 2006). Metamorphosis is likely to occur during the summer or autumn period (Anstis, 2002).

Litoria olongburensis is the most threatened ‘acid’ frog species, and is listed in QLD, N.S.W. and Australian legislation as Vulnerable (Table 1.1). In addition to those threatening processes outlined previously Meyer et al. (2006) indicates that the mosquito fish (Gambusia holbrooki) may threaten populations and that competition from L. fallax may occur in disturbed areas. Despite these threats, the amount of peer-reviewed literature on the biology and ecology of L. olongburensis is limited, particularly in relation to the non-breeding habitat requirements or factors that influence this species’ distribution (Hines et al., 1999; Meyer et al., 2006). Only a single publication provides data on habitat preference of L. olongburensis in north-eastern N.S.W. (Lewis and Goldingay, 2005).

Figure 1.2: The four Australian ‘acid’ frog species – 1. Litoria olongburensis; 2. Crinia tinnula; 3. Litoria freycineti; 4. Litoria cooloolensis 20

1.4.2.2 Crinia tinnula

The second study species, C. tinnula, belongs to the Myobatrachidae family and is often referred to as the Wallum Froglet or the Tinkling Frog (Meyer et al., 2006). The species can range between 16-22mm SVL and vary in coloration, with the dorsum of different individuals being recorded as beige, red or dark-brown with irregular markings or stripes (Figure 1.2) (Meyer et al., 2006). Ventral coloration is also variable and may be white with dark grey flecking, peppered grey with black and white or white flecking on dark grey (Meyer et al., 2006). There will usually be a pale stripe that runs from the throat through the middle of the belly (Meyer et al., 2006) and occasionally another stripe running from armpit to armpit (Robinson 2002).

The mainland distribution of C. tinnula occurs between Littabella National Park, QLD, to South Kurnell, Sydney, N.S.W. (Figure 1.4) in acidic Melaleuca (paperbark) and sedge swamps (Meyer et al., 2006). Records have also been located in disturbed sites (i.e. pine plantations, burnt heath) with sister species co-occurrence (i.e. Crinia signifera) (Simpkins, per. obs.)

Breeding occurs in spring, late summer, autumn and winter when males call (Anstis, 2002; Meyer et al., 2006). Tadpoles can grow to 36.7mm in tail length and 11.4mm in body length at Gosner stage 35 (Anstis, 2002). Tadpoles are usually found at the bottom of waterbodies (Anstis, 2002) in shallow waters (< 1m depth) (Meyer et al., 2006). Metamorphosis is likely to occur after about six months, however, this was observed in captivity over winter (Anstis, 2002).

Crinia tinnula is the second most threatened ‘acid’ frog species, being listed in both QLD and N.S.W. as vulnerable (Table 1.1). There is currently no peer-reviewed literature on non-breeding habitat requirements or factors that influence this species’ distribution (Hines et al., 1999).

Figure 1.3: Distribution of L. olongburensis as indicated by red and blue circles. Red circles indicate records obtained between 1995-2004.Blue circles indicate records obtained before 1995. Regional boundaries are indicated by grey lines. Dotted line represents the Queensland / New South Wales state boundary. Records sourced from EPA/QPWS, NSWDEC, the Australian Museum, Queensland Museum, South Australian Museum, and various biologists. Figure obtained from Meyer et al., 2006. 22

Figure 1.4: Distribution of C. tinnula as indicated by red and blue circles. Red circles indicate records obtained between 1995-2004.Blue circles indicate records obtained before 1995. Regional boundaries are indicated by grey lines. Dotted line represents the Queensland / New South Wales state boundary. Records sourced from EPA/QPWS, NSWDEC, the Australian Museum, Queensland Museum, South Australian Museum, and various biologists. Figure obtained from Meyer et al., 2006.

1.5 Study Aims

This study aimed to identify those environmental factors influencing the tadpole and adult amphibian assemblages (with primary focus on L. olongburensis) in the wallum heathland of eastern Australia. Hypotheses were derived from current information suggesting that fish (i.e. introduced species), competition and water chemistry (i.e. pH) influence the distribution and assemblage of these tadpoles and adult amphibians.

The first three chapters of this thesis focus on determining which environmental variables are important in structuring the adult and tadpole amphibian assemblage within and around wallum heathland areas. Furthermore, the first of these three chapters will focus on determining how biotic and abiotic variables are influencing the two most threatened ‘acid’ frog species, L. olongburensis and C. tinnula. The forth chapter aimed at determining predation rates of L. olongburensis and two other amphibian tadpole species to both native and exotic aquatic predators through an experimental manipulation.

The results from this study are urgently needed for conservation/management purposes as there is little peer-reviewed information on any wallum associated amphibian species. This knowledge will allow managers and scientists to evaluate suitable habitat for threatened species associated with wallum heathland, and prioritise areas for conservation. This research is also essential for environmental consultants performing Environmental Impact Assessments (EIA) to identify important breeding habitat for L. olongburensis and C. tinnula. Futhermore, the study will determine if anthropogenic habitats can be constructed in an attempt to mitigate against habitat loss when develeopment occurs in areas where L. olongburensis are present.

Each chapter has been written in draft format for publication purposes. Consequently, individual chapters have been formatted to meet the style requirements of target peer- reviewed journals.

1.6 References

Altwegg, R., Reyer, H.-U., 2003. Patterns of Natural Selection on Size at Metamorphosis in Water Frogs. Evolution 57 (4), 872-882. Álvarez, D., Nicieza, A. G., 2009. Differential success of prey escaping predators: tadpole vulnerability or predator selection? Copeia 2009, 453-457. Andrén, C., Henrikson, L., Olsson, M., Nilson, G., 1988. Effects of pH and aluminium on embryonic and early larval stages of Swedish brown frogs Rana arvalis, R. temporaria and R. dalmatina. . Ecography 11 (2), 127-135. Anstis, M. 2002. Tadpoles of South-eastern Australia: a guide with keys, New Holland Publishers, Sydney. Arendt, J. D., 2009. Infuence of sprint speed and body size on predator avo in New Mexican spadefoot toads (Spea multiplicata). Oecologia 159, 455-461. Babbitt, K. J., Baber, M. J., Tarr, T. L., 2003. Patterns of larval amphibian distribution along a wetland hydroperiod gradient. Canadian Journal of Zoology 1552, 1539- 1552. Babbitt, K. J., Tanner, G. W., 1998. Effects of cover and predator size on survival and development of Rana utricularia tadpoles. Oecologia 114, 258-262. Baber, M. J., Babbitt, K. J., 2003. The relative impacts of native and introduced predatory fish on a temporary wetland tadpole assemblage. Oecologia 136, 289-295. Baber, M. J., Fleishman, E., Babbitt, K. J., Tarr, T. L., 2004. The relationship between wetland hydroperiod and nestedness patterns in assemblages of larval amphibians and predatory macroinvertebrates. Oikos 107, 16-27. Barth, B. J., Wilson, R. S., 2010. Life in acid: interactive effects of pH and natural organic acids on growth, development and locomotor performance of larval striped marsh frogs (Limnodynastes peronii). The Journal of Experimental Biology 213, 1293- 1300. Beebee, T. J. C., Denton, J. S., Buckley, J., 1996. Factors affecting population densities of adult Natterjack Toads Bufo calamita in Britain. Journal of Applied Ecology 33, 263-268. Bentley, P.J., Yorio, T., 1979. Do frogs drink? Journal of Experimental Biology 79, 41-46. 25

Berven, K. A., 1990. Factors affecting population fluctuations in larval and Adult Stages of the Wood Frog (Rana sylvatica). Ecology 71, 1599-1608. Binckley, C. A., Resetarits Jr., W. J., 2003. Functional equivalence of non-lethal effects: generalized fish avoidance determines distribution of gray treefrog, Hyla chrysoscelis, larvae. Oikos 102, 623-629. Brodie III, E. D., Brodie Jr., E. D., 1999. Predator-Prey Arms Races. Bioscience 49, 557- 568. Brodie Jr., E. D., Formanowicz Jr., D. R., 1987. Antipredator mechanisms of larval anurans: protection of palatable individuals. Herpetologica 43, 369-373. Carey, C. (1978). Factors affecting body temperatures of toads. Oecologia 35, 197-219. Chinathamby, K., Reina, R. D., Bailey, P. C. E., Lees, B. K., 2006. Effects of salinity on the survival, growth and development of tadpoles of the brown tree frog, Litoria ewingii. Australian Journal of Zoology 54, 97-105. Christy, M. T., Dickman, C. R., 2002. Effects of salinity on tadpoles of the green and golden bell frog (Litoria aurea). Amphibia-Reptilia 23, 1-11. Coaldrake, J.E., 1961. The eco-system of the coastal lowlands (Wallum) of southern Queensland. C.S.I.R.O. Bulletin. No. 283. Crossland, M. R., Alfrod, R. A., Shine, R., 2009. Impact of the invasive cane toad (Bufo marinus) on an Australian frog (Opisthodon ornatus) depends on minor variation in reproductive timing. Oecologia 158, 625-632. Crump, M. L. (2010) Amphibian diversity and life history. In: Dodd Jr. CK (ed) Amphibian Ecology and Conservation: A Handbook of Techniques. Oxford University Press, New York. Cummings, C. P., 1986. Effects of aluminum and low pH on growth and development in Rana temporaria tadpoles. . Oecologia 69 (2), 248-252. Cushman, S., 2006. Effects of habitat loss and fragmentation on amphibians: A review and prospectus. Biological Conservation 128, 231-240. Death , R. G., Winterborn, M. J., 1995. Diversity patterns in stream benthic invertebrate communities: The influence of habitat stability. Ecology 76, 1446-1460. Duellmann, W. E., Trueb, L. 1986. Biology of amphibians, John Hopkins University Press, Baltimore. 26

Feder, M. E., 1983. The relation of air breathing and locomotion to predation on tadpoles, Rana berlandieri, by turtles. Physiological Zoology 56 (4), 522-531. Flecker, A. S., Feifarek, B. P., Taylor, B. D., 1999. Ecosystem engineering by a tropical tadpoles: density-dependant effects on habitat structure and larval growth rates. Copeia 1999 (2), 495-500. Fox, P. J., 1978. Caddis larvae (Trichoptera) as predators of fish eggs. Freshwater Biology 8, 343-345. Ferrari, M. C. O., Chivers, D.P., 2008. Temporal variability, threat sensitivity and conflicting information about the nature of risk: understanding the dynamics of tadpole antipredator behaviour. Animal Behaviour 78, 11-16. Freda, J., 1986. The influence of acidic pond water on amphibians: A review. Water, Air and Soil Pollution 30, 439-450. Freda, J., Taylor, D. H., 1992. Behavioral response of amphibian larvae to acidic water. Journal of Herpetology 26, 429-433. Gibbs, J. P., 1998. Distribution of woodland amphibians along a forest fragmentation gradient. Landscape Ecology, 263-268. Gillespie, G., Hero, J.-M. (1999) Potential impacts of introduced fish and fish translocations on Australian amphibians. In: Campbell A (ed) Declines and Disappearances of Australian Frogs. Environment Australia, Canberra, pp 131-144. Girish, K. G., Krishna-Murthy, S. V. B., 2009. Distribution of tadpoles of large wrinkled frog Nyctibatrachus major in central Western Ghats: influence of habitat variables. Acta Herpetologica 4, 153-160. Glos, J., Grafe, T.U., Rödel, M.-O., Linsenmair, K.W., 2003. Geographic variation in pH tolerance of two populations of the European common frog, Rana temporaria. Copeia 2003, 650-656. Goldberg, F. J., Quinzio, S., Vaira, M., 2006. Oviposition-site selection by the toad Melanophryniscus rubriventris in an unpredictable environment in Argentina. Canadian Journal of Zoology 84, 699-705. Gomez-Mestre, I., Tejedo, M., Ramayo, E., Estepa, J., 2004. Developmental Alterations and osmoregulatory physiology of a larval anuran under osmotic stress. Physiological and Biochemical Zoology 77, 267-274. 27

Gosner, K. L., Black, I. H., 1957. The effects of acidity on the development and hatching of New Jersey Frogs. Ecology 38, 256-262. Grand, J., Cushman, S. A., 2003. A multiple-scale analysis of species-habitat relationships: breeding birds in a pitch pine-scrub oak (Pinus rigida-Quercus ilicifolia) community. Biological Conservation 112 (3), 307-317. Gregoire, D. R., Gunzburger, M. S., 2008. Effects of predatory fish on survival and behavior of Larval Gopher Frogs (Rana capito) and Southern Leopard Frogs (Rana sphenocephala). Journal of Herpetology 42, 97-103. Griffith, S. J., Bale, C., Adam, P., 2008. Environmental correlates of coastal heathland and allied vegetation. Australian Journal of Botany 56, 512-526. Griffith, S. J., Bale, C., Adam, P., Wilson, R., 2003. Wallum and related vegetation on the NSW North Coast: description and phytosociological analysis. Cunninghamia 8 (2), 202-252. Griffith, S.J., Bale, C., Adam, P., 2004. The influence of fire and rainfall on seedling recruitment in sand mass (wallum) heathland of north-eastern New South Wales. Australian Journal of Botany 52, 93-118. Griffiths, R. A., Denton, J., Wong, A. L.-C., 1993. The effect of food level on competition in tadpoles: interference mediated by Protothecan algae? Journal of Animal Ecology 62, 274-279. Groves, R. H., 1981. Heathland soils and their fertility status. In Specht, R.L. Ecosystems of the world 9B: heathlands and related shrublands; analytical studies. Elservier Scientific Publishing Company, Amsterdam. Pp143-149. Gunzburger, M. S., Travis, J., 2005. Critical literature review of the evidence of unpalatability of amphibian eggs and larvae. Journal of Herpetology 39, 547-571. Haddad, C. F. B., Prado, C. P. A., 2005. Reproductive modes in frogs and their unexpected diversity in the Atlantic forest in Brazil. Bioscience Biotechnology and Biochemistry 55 (3), Halliday, T. R., 2008. Why amphibians are important. International Zoo Yearbook 42, 7- 14. Haramura, T., 2008. Experimental test of spawning site selection by Buergeria japonica (Anura: Rhacophoridae) in response to salinity level. Copeia 2008, 64-67. 28

Hecnar, S. J., M'Closkey, R. T., 1997. The effects of predatory fish on amphibian species richness and distribution. Biological Conservation 79, 123-131. Henrikson, B. I., 1990. Predation on amphibian eggs and tadpoles by common predators in acidified lakes. Holartic Ecology 13, 201-206. Hero, J.-M., Gascon, C., Magnusson, W. E., 1998. Direct and indirect effects of predation on tadpole community structure in the Amazon rainforest. Australian Journal of Ecology 23, 474-482. Hero, J.-M., Magnusson, W. E., Rocha, C. F. D., Catterall, C. P., 2001. Antipredator defenses influence the distribution of amphibian prey species in the central Amazon rain forest. Biotropica 33, 131-141. Hero, J.-M., Morrison, C., 2004. Frog declines in Australia: Global implications. Herpetological Journal 14 (4), 175-186. Hero, J.-M., Morrison, C., Gillespie, G., Roberts, J.D., Newell, D., Meyer, E., McDonald, K., Lemckert, F., Mahony, M., Osborne, W., Hines, H., Richards, S., Hoskin, C., Clarke, J., Doak, N. and Shoo, L. (2006). Overview of the conservation status of Australian Frogs. Pacific Conservation Biology 12, 313-320. Herrmann, H. L., Babbitt, K. J., Baber, M. J., Congalton, R. G., 2005. Effects of landscape characteristics on amphibian distribution in a forest-dominated landscape. Biological Conservation 123, 139-149. Heyer, W. R., 1973. Ecological Interactions of Frog Larvae at a Seasonal Tropical Location in Thailand. Journal of Herpetology 7, 337-361. Heyer, W. R., McDiarmid, R. W., Weigmann, D. L., 1975. Tadpoles, predation and pond habitats in the tropics. Biotropica 7, 100-111. Hickman Jr., C. P., Roberts, L. S., Larson, A. 1998. Biology of amphibians, McGraw-Hill Companies, United States of America. Hines, H., Mahony, M., McDonald, K. (1999) An assessment of frog declines in wet subtropical Australia. In: Campbell A (ed) Declines and Disappearances of Australian Frogs. Environment Australia, Canberra, pp 44-63. Hossie, T. J., Murray, D. L., 2010. You can't run but you can hide: refuge use in frog tadpoles elicits density-dependent predation by dragonfly larvae. Oecologia 163, 395-404. 29

Ingram, G. J., Corben, C. J., 1975. The frog fauna of North Stradbroke Island, with comments on the 'acid' frogs of the wallum. The Proceedings of the Royal Society of Queensland 86, 49-54. Jara, F. G., 2008. Tadpole–odonate larvae interactions: influence of body size and diel rhythm. Aquatic Ecology 42, 503-509. Jara, F. G., Perotti, M. G., 2008. Toad tadpole responses to predator risk: ontogenetic change between constitutive and inducible defenses. Journal of Herpetology 43 (1), 82-88. Johnson, J. R., Knouft, J. H., Semlitsch, R. D., 2007. Sex and seasonal differences in the spatial terrestrial distribution of gray treefrog (Hyla versicolor) populations. Biological Conservation 140, 250-258. Kats, L. B., Ferrer, R. P., 2003. Alien predators and amphibian declines: review of two decades of science and the transition to conservation. Diversity and Distributions 2003, 99-110. Kats, L. B., Petranka, J. W., Sih, A., 1988. Antipredator defenses and the persistence of amphibian larvae with fishes. Ecology 69, 1865-1870. Kiesecker, J. M., Blausien, A. R., Swihart, R. K., 2001. Complex causes of amphibian population declines. Nature 410, 681-684. Kolozsvary, M. B., Swihart, R. K., 1999. Habitat fragmentation and the distribution of amphibians: patch and landscape correlates in farmland. Canadian Journal of Zoology 77, Kopp, K., Wachlevski, M., Eterovick, P. C., 2006. Environmental complexity reduces tadpole predation by water bugs. Canadian Journal of Zoology 84, 136-140. Lane, M. J., Mahony, S. J., 2002. Larval anurans with synchronous and asynchronous development periods: contrasting responses to water reduction and predator presence. Journal of Animal Ecology 71, 780-792. Lawler, K. L., Hero, J.-M., 1997. Palatability of Bufo marinus tadpoles to a predatory fish decreases with development. Wildlife Research 24, 327-334. Lemckert, L., Haywood, A., Brassil, T., Mahony, M. (2006). Correlations between frogs and pond attributes in central New South Wales, Australia: What makes a good pond? Applied Herpetology 3, 67-81. 30

Lewis, B. D., Goldingay, R. L., 2005. Population monitoring of the vulnerable wallum sedge frog (Litoria olongburensis) in north-eastern New South Wales. Australian Journal of Zoology 53, 185-194. Loman, J., 1999. Early metamorphosis in common frog Rana temporaria tadpoles at risk of drying: an experimental demonstration. Amphibia-Reptilia 20, 421-430. Loman, J., Claesson, D., 2003. Plastic response to pond drying in tadpoles Rana temporaria: test of cost models. Evolutionary Ecology and Research 5, 179-194. Lowe, K., Hero, J.-M., 2012. Sexual dimorphism and color polymorphism in the Wallum Sedge Frog (Litoria olongburensis). Herpetological Review 43, 236-240. Mann, R., Bidwell, J. (1999) Toxicological issues for amphibians in Australia. In: Campbell A (ed) Declines and Disappearances of Australian Frogs. Environment Australia, Canberra, pp 185-201 Márquez-García, M., Correa-Solís, M., Méndez, M. A., 2010. Life-history trait variation in tadpoles of the warty toad in response to pond drying. Journal of Zoology 281, 105- 111. Marsh, D.M., Fegraus, E.H. and Harrison, S. (1999). Effects of breeding pond isoloation on spatial and temporal dynamics of pond use by the tungara frog, Physalaemus pustulosus. Journal of Animal Ecology 68, 804-814. Marsh, D.M. and Trenham, P.C. (2001). Metapopulation dynamics and amphibian conservation. Conservation Biology 15, 40-49. Mazerolle, M. J., 2005. Peatlands and green frogs: A relationship regulated by acidity? Ecoscience 12, 60-67. McCollum, S. A., Leimberger, J. D., 1997. Predator-induced morphological changes in an amphibian: predation by dragonflies affects tadpole shape and color. Oecologia 109, 615-621. McDiarmid, R. W., Altig, R. (2010) Morphology of amphibian larvae. In: Dodd Jr. CK (ed) Amphibian Ecology and Conservation: A Handbook of Techniques. Oxford, New York, pp 39-53 Meyer, E. 2004. Acid adaptation and mechanisms for softwater acid tolerance in larvae of anuran species native to the 'Wallum' of east Australia. PhD Thesis. University of Queensland. 31

Meyer, E., Hero, J.-M., Shoo, L., Lewis, B. (2006) National recovery plan for the wallum sedgefrog and other wallum-dependent frog species. Report to the Department of the Environment and Water Resources, Canberra. Queensland Parks and Wildlife Service, Brisbane, Michel, M. J., Adams, M. M., 2009. Differential effects of structural complexity on predator foraging behavior. Behavioral Ecology 20, 313-317. Mokany, A., 2007. Impacts of tadpoles and mosquito larvae on ephemeral pond structure and processes. Marine and Freshwater Research 20 (2), 313-317. Mokany, A., Shine, R., 2003. Competition between tadpoles and mosquito larvae. Oecologia 135, 615-620. Moreira, L. F. B., Machado, I. F., Garcia, T. V., Maltchik, L., 2010. Factors influencing anuran distribution in coastal dune wetlands in southern Brazil. Journal of Natural History 44, 1493-1507. Morin, P. J., 1986. Interactions between intraspecific competition and predation in an amphibian predator-prey system. Ecology 67 (3), 713-720. Morin, P. J., 1995. Functional redundancy, non-additive interactions, and supply-side dynamics in experimental pond communities. Ecology 76, 133-149. Newman, R. A., 1994. Effects of Changing Density and Food Level on Metamorphosis of a Desert Amphibian, Scaphiopus couchii. . Ecology 75 (4), 1085-1096. Noland, R., Ultsch, G. R., 1981. The roles of temperature and dissolved oxygen in micorhabitat selection by the tadpoles of a frog (Rana pipiens) and a toad (Bufo terrestris). Copiea 1981 (3), 645-652. Orizaola, G., BraÑa, F., 2003. Do predator chemical cues affect oviposition site selection in newts? Herpetological Journal 13, 189-193. Orizaola, G., Laurila, A., 2009. Intraspecific variation of temperature-induced effects on metamorphosis in the pool frog (Rana lessonae). Canadian Journal of Zoology 87, 581-588. Oseen, K. L., Wassersug, R. J., 2002. Environmental factors influencing calling in sympatric anurans. Oecologia 133, 616-625. Pearson, O.P. and Bradford, D.F. (1976). Thermoregulation of lizards and toads at high altitudes in southern Peru. Copeia 176, 155-170. 32

Persson, M., Räsänen, K., Laurila, A., Merilä, J., 2007. Maternally determined adaptation to acidity in Rana arvalis: Are laboratory and field estimates of embryonic stress tolerance congruent? Canadian Journal of Zoology 85, 832-838. Pfennig, D. W., Murphy, P. J., 2002. How fluctuating competition and phenotypic plasticity mediate species divergence. Evolution 56, 1217-1228. Picker, M. D., McKenzie, C. J., Fielding, P., 1993. Embryonic Tolerance of Xenopus (Anura) to Acidic Blackwater. Copeia 4, 1072-1081. Pierce, B. A., 1985. Acid tolerance in amphibians. Bioscience 35, 239-243. Pinder, A. W., Storey, K. B., Ultsch, G. R. (1992) Estivation and hibernation. In: Feder ME and Burggren, W.W (ed) Environmental physiology of the amphibia. Chicago Press, Chicago, pp 250-276 Polo-Cavia, N., Gonzalo, A., López, P., Martín, J., 2010. Predator recognition of native but not invasive turtle predators by naïve anuran tadpoles. Animal Behaviour 80, 461- 466. Ranvestal, A. W., Lips, K. R., Pringle, C. M., Whiles, M. R., Bixby, R. J., 2004. Neotropical tadpoles influence stream benthos: evidence for the ecological consequences of decline in amphibian populations. Freshwater Biology 49, 274- 285. Räsänen, K., Söderman, F., Laurila, A., Merilä, J., 2008. Geographic variation in maternal investment: acidity affects egg size and fecundity in Rana arvalis. Ecology 89, 2553-2562. Relyea, R. a., 2003. Predators Come and Predators Go: the Reversibility of Predator- Induced Traits. Ecology 84, 1840-1848. Relyea, R. A., 2007. Getting out alive: how predators affect the decision to metamorphose. Oecologia 152, 389-400. Resetarits Jr, W. J., Wilbur, H. M., 1989. Choice of oviposition site by Hyla chrysoscelis: role of predators and competitors. Ecology 70, 220-228. Retallick, R. W. R. 2000. Determinants of the assemblage structure of tadpoles in the streams of Eungella National Park. James Cook University Richards, S. J., 2002. Influence of flow regime on habitat selection by tadpoles in an Australian rainforest stream. Journal of Zoology 257, 273-279. 33

Richter-Boix, A., Llorente, G. A., Montori, A., 2007. A comparative study of predator- induced phenotype in tadpoles across a pond permanency gradient. Hydrobiologia 583, 43-56. Rios-López, N., 2008. Effects of increased salinity on tadpoles of two anurans from a Caribbean coastal wetland in relation to their natural abundance. Amphibia-Reptilia 29, 7-18. Robinson, M., 2002. A field guide to frogs of Australia: from Port Augusta to Fraser Island including Tasmania. Reed New Holland, Australia. Rödel, M.-O., Linsenmair, K. E., 1997. Predator-induced swarms in the tadpoles of an African savanna frog, Phyzomuntis mimps. Ethology 103, 902-914. Rudolf, V. H. W., Rödel, M.-O., 2007. Phenotypic plasticity and optimal timing of metamorphosis under uncertain time constraints. Evolutionary Ecology 21, 121- 142. Sadinski, W. A., Dunson, W. J., 1992. A multilevel study of effects of low pH on amphibians of temporary ponds. Journal of Herpetology 26, 413-422. Saidapur, S. K., Veeranagoudar, D. K., Hiragond, N. C., Shanbhag, B. A., 2009. Mechanism of predator–prey detection and behavioral responses in some anuran tadpoles. Chemoecology 19, 21-28. Schlichter, L., 1981. Low pH affects the fertilization and development of Rana pipiens eggs. Canadian Journal of Zoology 59, 1693-1699. Schmutzer, A. C., Gray, M. J., Burton, E. C., Miller, D. L., 2008. Impacts of cattle on amphibian larvae and the aquatic environment. Freshwater Biology 53, 2613-2625. Schoener, T. W., 1983. Field experiments on interspecific competition. The American Naturalist 122, 240-285. Scott, D. E., 1994. The effect of larval density on adult demographic traits in Ambystoma opacum. Ecology 75, 1383-1396. Semlitsch, R. D., Ryer, H. U., 1992. Performance of tadpoles from the hybridogenetic Rana esculenta complex: interactions with pond drying and interspecific competition. Evolution 46 (3), 665-676. Seymour, R. S., Roberts, J. D., Mitchell, N. J., Blaylock, A. J., 2000. Influence of environmental oxygen on development and hatching of aquatic eggs of the 34

Australian frog, Crinia georgiana. Physiological and Biochemical Zoology 73 (4), 501-507. Shah, A.A., Ryan, M.J., Bevilacqua, E., Schlaepfer, M.A., 2010. Prior experience alters the behavioral response of prey to a nonnative predator. Journal of Herpetology 44, 185-192. Shoemaker, V.H., Nagy, K.A., 1977. Osmoregulation in amphibians and reptiles. Annual Review of Physiology 1977, 449-471 Simpkins, C. A., Meyer, E., Hero, J.-M., 2011. Long-range movement in the rare Cooloola sedgefrog Litoria cooloolensis. Australian Zoologist 35, 977-978. Skelly, D. K., 1994. Activity level and susceptibility of anuran larvae to predation. Animal Behaviour 47, 465-468. Skelly, D. K., 1995. A Behavioral Trade-Off and Its Consequences for the Distribution of Pseudacris Treefrog Larvae. Ecology 76, 150-164. Smallbone, LT., Luck, G.W. and Wassens, S. (2011). Anuran species in urban landscapes: relationships with biophysical, built environment ans socio-economic factors. Landscape and Urban Palnning 101, 43-51. Smith, A. R., Awan, G. R., 2009. The roles of predator identity and group size in the antipredator responses of American toad (Bufo americanus) and bullfrog (Rana catesbeiana) tadpoles. Behaviour 146, 225-243. Smith, D. C., 1987. Adult recruitment in chorus frogs: effects of size and date at metemorphosis. Ecology 68, 344-350. Smith, G. R., Burgett, A. A., Temple, K. G., Sparks, K. A., Winter, K. E., 2008. The ability of three species of tadpoles to differentiate among potential fish predators. Ethology 114, 701-710. Smith, M. J., Schreiber, E. S. G., Scroggie, M. P., Kohout, M., Ough, K., Potts, J., Lennie, R., Turnbill, D., Jin, C., Clancy, T., 2007. Associations between anuran tadpoles and salinity in a landscape mosaic of wetlands impacted by secondary salinisation. Freshwater Biology 52, 75-84. Snodgress, J. W., Komoroski, M. J., Bryan Jr., A. L., Burger, J., 2000. Relationships among isolated wetland size, hydroperiod, and amphibian species richness: implications for wetland regulations. Conservation Biology 14, 414-419. 35

Sparling, D. W. (2010) Water-quality criteria for amphibians. In: Dodd Jr. CK (ed) Amphibian Ecology and Conservation: A Handbook of Techniques. Oxford University Press, Oxford, pp 105-120. Spieler, M., 2003. Risk of predation affects aggregation size: a study with tadpoles of Phrynomantis microps (Anura: Microhylidae). Animal Behaviour 65, 179-184. Specht, R.L., 1981. Conservation: Australian Heathlands. In Specht, R.L. Ecosystems of the World 9B: Heathlands and Related Shrublands; Analytical Studies. Elsevier Scientific Publishing Company, Amsterdam, pg 235-240. Stav, G., Kotler, B. P., Blaustein, L., 2007. Direct and indirect effects of dragonfly (Anax imperator) nymphs on green toad (Bufo viridis) tadpoles. Hydrobiologia 579, 85-93. Steinberg, C. E. W., Kamara, S., Prokhotskaya, V. Y., Manusadzianas, L., Karasyova, T. A., Timofeyev, M. A., Jie, Z., Paul, A., Meinelt, T., Farjalla, V. F., Matsuo, A. Y. O., Burnison, B. K., Menzel, R., 2006. Dissolved humic substances – ecological driving forces from the individual to the ecosystem level? Freshwater Biology 51, 1189-1210. Stoneham, M. 2001. The influence of stream-dwelling predators on the distribution and density of Mixophyes tadpoles in Southeast Queensland. Unpublished Honours Thesis, Griffith University. Strahan, R., 1957. The Effect of Salinity on the Survival of Larvae of Bufo melanostictus Schneider. Copeia 1957, 146-147. Stuart, S. N., Chanson, J. S., Cox, N. A., Young, B. E., Rodrigues, A. S. L., Fischman, D. L., Waller, R. W., 2004. Status and trends of amphibian declines and extinctions worldwide. Science 306, 1783-1786. Tarr, T. L., Babbitt, K. J., 2002. Effects of habitat complexity and predator identity on predation of Rana clamitans larvae. Amphibia-Reptilia 23, 13-20. Timofeyev, M. A., Shatilina, Z. M., Bedulina, D. S., Menzel, R., Steinberg, C. E. W., 2006. Natural organic matter (NOM) has the potential to modify and multixenobiotic resistance (MXR) activity in freshwater amphipods Eulimnogammarus cyaneus and E. verrucosus. Comparative Biochemistry and Physiology, Part B 146, 496-503. Toft, C. A., 1985. Resource Partitioning in Amphibians and Reptiles. Copeia 1985 (1), 1- 21. 36

Touchon, J. C., Warkentin, K. M., 2008. Fish and dragonfly nymph predators induce opposite shifts in color and morphology of tadpoles. Oikos 117, 634-640. Twomey, E., Morales, V., Summers, K., 2008. Evaluating condition-specific and asymmetric competition in a species-distribution context. Oikos 117, 1175-1184. Tyler, M. J., 1997. The Action Plan for Australian Frogs. Widlife Australia Canberra, ACT. Van Buskirk, J., Aschwanden, J., Buckelmüller, I., Reolon, S., Rüttiman, S., 2004. Bold tail coloration protects tadpole from dragonfly strikes. 2004, 599-602. Vonesh, J. R., Kraus, J. M., Rosenberg, J. S., Chase, J. M., 2009. Predator effects on aquatic community assembly: disentangling the roles of habitat selection and post- colonization processes. Oikos 118, 1219-1229. Wassersug, R. J., Seibert, E. A., 1975. Behavioral responses of amphibian larvae to variation in dissolved oxygen. Copeia 1975 (1), 86-103. Watt, P. J., Nottingham, S. F., Young, S., 1997. Toad tadpole aggregation behaviour: evidence for a predator avoidance function. Animal Behaviour 54, 865-872. Welborn, G. A., Skelly, D. K., Werner, E. E., 1996. Mechanisms creating community structure across a freshwater habitat gradient. Annual Review of Ecology and Systematics 27, 337-363. Wells, K. D. 2007. The Ecology and Behavior of Amphibians, The University of Chicago Press, USA Wilbur, H. M., 1987. Regulation of structure in complex systems: experimental temporary pond communities. Ecology 68, 1437-1452. Wilbur, H. M., 1997. Experimental ecology of food webs: compleex systems in temporary ponds. Ecology 78, 2279-2302. Wiltshire, D. J., Bull, C. M., 1977. Potential competitive interactions between larvae of Pseudophryne bibroni and P. semimarmorata (Anura: Leptodactylidae). Australian Journal of Zoology 25, 449-454. Wood, C. M., Matsuo, A. Y. O., Wilson, R. W., Gonzalez, R. J., Patrick, M. L., Playle, R. C., Val, A. L., 2003. Protection by natural blackwater against disturbances in ion fluxes caused by low pH exposure in freshwater stingrays endemic to the Rio Negro. Physiological and Biochemical Zoology 76, 12-27.

Woodward, B. D., 1983. Predator-prey interactions and breeding-pond use of temporary- pond species in a desert anuran community. Ecology 64, 1549-1555.

Chapter 2 - Environmental variables associated with the distribution and occupancy of tadpoles in naturally acidic, oligotrophic waterbodies

2.1 Abstract

Environmental factors play an integral role, either directly or indirectly, in structuring faunal assemblages. Water chemistry, predation, hydroperiod and competition influence tadpole assemblages within waterbodies. I surveyed aquatic predators, habitat refugia, water height and water chemistry variables (pH, salinity and turbidity) at 37 waterbodies over an intensive 22 day field survey to determine which environmental factors influence the relative abundance and occupancy of anuran tadpole species in naturally acidic, oligotrophic waterbodies within eastern Australian wallum communities. The majority of tadpoles found were of Litoria olongburensis (Wallum Sedge Frog) and Crinia tinnula (Wallum Froglet) species, both habitat specialists that are associated with wallum waterbodies and listed as Vulnerable under the IUCN Red List. Tadpoles of two other species (Litoria fallax (Eastern Sedge Frog), and Litoria cooloolensis (Cooloola Sedge Frog)) were recorded from two waterbodies. Tadpoles of Litoria gracilenta (Graceful Treefrog) were recorded from one waterbody. Relative abundance and occupancy of L. olongburensis tadpoles were associated with pH and water depth. Additionally, L. olongburensis tadpole relative abundance was negatively associated with turbidity. Waterbody occupancy by C. tinnula tadpoles was negatively associated with predatory fish and water depth and positively associated with turbidity. Variables associated with relative abundance of C. tinnula tadpoles were inconclusive and further survey work is required to identify these environmental factors. These results show that the ecology of tadpole species associated with wallum waterbodies is complex and species specific. Therefore, several environmental factors require consideration for successful management of waterbodies where the conservation of threatened wallum amphibian communities is a priority. These results may also be relevant in assisting scientists and managers in determining how

environmental variables are influencing tadpole distributions for species that are associated with naturally acidic, oligotrophic waters around the globe.

2.2 Introduction

Environmental factors play an integral role, either directly or indirectly, in structuring faunal assemblages (Krebs 2009). Different species will have varying tolerances to environmental factors (e.g. reviewed in Pierce 1985; Berkelmans & Willis 1999; Schofield & Nico 2009), with organisms occurring outside of their tolerance range often resulting in death or deformities (Gosner & Black 1957; Sadinski & Dunson 1992; Schofield & Nico 2009). Therefore, organisms only occur in areas where environmental variables are within their tolerance limits.

The anuran tadpole is the non-reproductive larval stage that occurs in different ecological niches to the adult stage (Wells 2007) and is critical in the amphibian lifecycle with ‘the potential to have the greatest impact on the continuing persistence of a population’ (Lane & Mahony 2002). It is imperative for amphibian conservation to determine which environmental factors influence individual abundance and occupancy of species within amphibian larval assemblages. Several factors may structure tadpole assemblages, including: intolerances to water chemistry (Smith et al. 2007; Wells 2007), hydroperiod (Snodgrass et al. 2000; Baber et al. 2004; Moreira et al. 2010), predation (Hero et al. 1998; Hero et al. 2001; Vonesh et al. 2009) and competition (Wiltshire & Bull 1977; Twomey et al. 2008).

Water chemistry and hydroperiod can influence amphibian assemblages, with numerous studies focusing on either water chemistry tolerance limits under laboratory conditions or water chemistry variables influencing amphibian assemblages. These variables include pH/acidity (reviewed in Freda 1986; Freda & Taylor 1992; Persson et al. 2007; Barth & Wilson 2010) salinity (Christy & Dickman 2002; Chinathamby et al. 2006; Rios-López 2008) and turbidity, as represented by suspended sediment (Knutson et al. 2004). Hydroperiod can structure the tadpole assemblage by excluding amphibian species with 40

long-lived tadpole stages, as these species can fail to metamorphose before desiccation occurs from temporary waterbodies (Welborn et al. 1996)..

Aquatic predators have also been suggested as a primary factor structuring tadpole assemblages (Heyer et al. 1975; Hero et al. 2001; Vonesh et al. 2009). The risk of predation can be determined by a number of factors, including the tadpole’s vulnerability to the predator assemblage or ability to avoid predators (i.e. via refugia usage) (Kopp et al. 2006; Saidapur et al. 2009). Competition may also structure tadpole communities by altering rates of tadpole growth and development (Wilbur 1987; Mokany & Shine 2002; Twomey et al. 2008), potentially leading to prolonged exposure to aquatic predators.

The coastal wallum vegetation along the eastern coast of Australia is the primary habitat for two frog species (Litoria olongburensis and Crinia tinnula) that are listed as Vulnerable under the IUCN Red List (IUCN 2011). Both species are referred to as ‘acid’ frog due to their association with low pH waters (Ingram and Corben 1975). For the purposes of this study ‘wallum’ will be referred to as the vegetation communities within the study area including Banksia woodland, sedgeland, heathland and Melaleuca swamps (Hines et al 1999, Griffith et al. 2003) that contain soils that are often acidic (pH 3.4-5.4) and low in nitrogen (Griffith et al. 2008) and phosphorus (Groves 1981). ‘Acid’ frog populations within protected areas are believed to be stable (Hines et al. 1999; Lewis and Goldingay 2005). However, populations of ‘acid’ frogs occurring outside of protected areas are at risk from habitat loss (Hines et al. 1999), with more than 30% of original heathland and Melaleuca cover in south-east Queensland cleared between 1974-1989 (Catterall and Kingston 1993, cited in Hines et al. 1999). It is therefore vital that mangers know which environmental factors influence ‘acid’ frogs within protected areas to ensure these environmental variables remain constant and populations remain stable. This knowledge would also aid managers, scientists and environmental consultants when evaluating suitable waterbodies for successful breeding of L. olongburensis and C.tinnula and, thus, be able to prioritize areas for conservation with greater accuracy. However, surveys to determine factors influencing L. olongburensis and C. tinnula tadpole relative abundance and occupancy within protected 41

wallum heathland waterbodies have never been published in peer-reviewed journals. Despite this, it has been suggested that pH (Ingram & Corben 1975) and introduced predatory fish (i.e. Gambusia holbrooki) (Meyer et al. 2006) are two important factors that may influence tadpole assemblages within these waterbodies.

I therefore investigated the influence of water chemistry factors (pH, salinity, turbidity), water depth, aquatic predators (predatory fish and aquatic invertebrates), competitors (tadpoles of other species) and tadpole refuge availability (represented by habitat complexity), on the occupancy and relative abundance of L. olongburensis and C. tinnula tadpoles across available habitats within protected areas of eastern Australia. I also examine the optimal ranges for factors that were observed to influence the relative abundance of L. olongburensis tadpoles.

2.3 Methods

Study Site Selection and Sampling Design

Nine national parks and reserves where L. olongburensis were known to occur were selected between the Great Sandy National Park, Queensland (QLD) (26.014° S, 153.024° E), and Wooli, New South Wales (NSW) (29.853° S, 153.263° E), in coastal wallum vegetation communities on mainland Australia. Eleven ‘survey transects’ were established across the national parks and reserves (one at each national park or reserve for all areas except Bundjalung National Park and Yuraygir National Park which contained two survey transect lines) to select waterbodies (Figure 2.1). With the exception of areas where no national parks or protected areas occur (i.e. major cities), the survey transects were established evenly across the known distributional range of L. olongburensis (Meyer et al. 2006) (Figure 2.1). Wallum vegetation communities often consist of several vegetation habitats (Ingram & Corben 1975, Hines et al. 1999). Therefore, survey transects were placed using QLD regional ecosystem and NSW vegetation G.I.S. layers to ensure all vegetation communities were intersected at least once from one of the survey transects. Additionally, historical locality records of L. olongburensis were used to ensure each 42

survey transect intersected a minimum of one waterbody where L. olongburensis had been recorded. Survey transects spanned from one perimeter of the protected area to the opposite perimeter and, therefore, varied in length. Anthropogenic disturbances (i.e. urban development) outside of national parks may influence environmental factors like water depth/hydrology, increase in nutrients or change in water chemistry (Meyer et al. 2006). Therefore, national parks and reserves were selected for surveying in an attempt to minimize the influence of anthropogenic disturbances, which are apparent throughout the range of L. olongburensis (Meyer et al. 2006).

Survey transects were then uploaded onto a Trimble® Juno™ SC unit for ground-based surveys targeting waterbodies. A maximum of two days was allowed for surveying of each survey transect.

Waterbodies were sampled if the survey transect intersected the waterbody or if the waterbody could be seen from the survey transect. Each waterbody had a ‘waterbody transect’ that was perpendicular to the edge of the waterbody, and went from one edge of the waterbody to the opposite edge. Where practical, the waterbody transect was placed through the deepest section of the waterbody. Waterbody transects varied in length depending on the waterbody size, with the shortest and longest waterbody transect being 48 meters and 200 meters, respectively.

Sites with large waterbodies containing heterogeneous vegetation had multiple waterbody transects to capture this variation. In these sites, primarily Bundjalung National Park and Yuraygir National Park, separate transects were placed a minimum of 10m apart in each distinct waterbody vegetation type to reduce edge effects and maintain independence.

All waterbody transects were surveyed over a 22 day period, in March 2010, to minimize the influence of temporal variation. This time of year was chosen as peak activity calling for L. olongburensis was predicted to occur in February and March (Hopkins unpublished data) and it was assumed that tadpoles could be present from eggs deposited in February.

This time also coincided with another survey that aimed at determining what environmental factors influenced the relative abundance of L. olongburensis adults (Shuker 2012).

It must be acknowledged that, due to the timing of the survey, winter breeding species may have been underepresented or missed during this study.

Dipnetting of aquatic fauna and measurements of water and vegetation characteristics were undertaken at five evenly distributed sampling points along the waterbody transects. Sampling point placement began 5 meters from the waterbody transect edge. Diurnal dipnetting was conducted using a circular net, with an aperture of approximately 30 centimetres (cm) in diameter and with mesh size less than 0.5 millimetres to capture tadpoles and aquatic predators. In an attempt to sample the entire water column three water column levels (bottom, middle and top) were dipnetted, with five sweeps at each level. To ensure no recaptures occurred animals were not released until dip-netting had been completed at each sampling point. It was assumed sampling points were sufficiently spaced (minimum distance between sampling points was 9.5 m) to prevent recaptures between successive sampling points.

Waterbody pH, salinity, turbidity, depth (as a representation of hydroperiod) and percent vegetation cover were also measured. A handheld TPS Aqua-CPA Conductivity-TDS-pH- Temperature Meter (version 1.2) was used to measure pH and salinity. Calibration occurred between survey transects to ensure accurate measurements. A 2 point calibration curve was used at 4.00 and 6.88 for pH calibration, while 2.00 parts per thousand (ppt) and 0.00 ppt were used for salinity calibration. Turbidity was measured by placing a black and white marker into the bottom of a transparent turbidity tube. Water was then added to the turbidity tube until the black and white marker could not be seen when looking directly down into the tube. The height of the water within the turbidity tube was then correlated to the turbidity value for the waters height. Water for turbidity measurements was sampled approximately 10 cm below the water’s surface. Water was sampled as close to the waters surface as possible when water depth fell below 10cm. Water depth was measured to the nearest centimetre in the centre of each sampling point. I compare waterbody size using the 45

waterbody transect lengths due to inadequate detail in available GIS layers, the size of the waterbodies and dense vegetation making it impractical to measure waterbody area during the surveys.

Aquatic sedge and herb cover was visually estimated using an estimated 5m x 5m quadrate centered on each sampling point. Percent cover was combined for each waterbody transect and divided by the number of sampling points (5) to give representation of percent cover for each waterbody. Aquatic sedge and herb cover was defined as the percentage of the quadrate occupied by the vertical projection of foliage and was used as an estimate of habitat complexity.

Rainfall data between December 2009 – March 2010 was obtained from the Australian Bureau of Meteorology Weather Stations located at Double Island Point Lighthouse (Weather Station # 040068), Beerburrum Forest Reserve (Weather Station # 040284), Byron Bay Cap Byron Lighthouse (Weather Station # 058009) and Wooli Beach (Weather Station # 058080). The rainfall averages for each month were obtained for all years that the weather stations had been in service.

Statistical Analysis

Water depth, water chemistry factors and all species of aquatic predators obtained from dipnetting were combined for each waterbody transect. One waterbody located next to an estuarine creek system contained an extreme mean salinity level of 420 parts per million and had recently been disturbed by fire, and was removed from the analysis. Water depth and water chemistry factors were divided by the number of sampling points (5) to give a mean per waterbody transect for each water chemistry variable. Tadpole species were considered present from a waterbody transect if a species was recorded from one of the five sampling points. Tadpole numbers were added together for each waterbody transect for each species for relative abundance analysis. Tadpoles used for analysis were of Gosner Stage 25 or later due to difficulties identifying tadpoles in earlier Gosner Stages. Fish from the species Gambusia holbrooki (Mosquito Fish), Rhadinocentrus ornatus (Ornate 46

Rainbow Fish), Hypseleotris galii (Firetail Gudgeon) and Nannoperca oxleyana (Oxleyan Pygmy Perch) were grouped into the ‘predatory fish’ category to estimate relative fish abundance for each waterbody transect. Predatory aquatic invertebrates (Belostomatidae (Giant Water Bugs) and Aeshnidae (Dragonfly families) were excluded from analysis as they were only detected from two transects.

To determine which variables were correlated a Spearman Rank Correlation Test was performed in IBM SPSS Statistics Version 19 (SPSS, Inc., 2009, Chicago). Variables that had a correlation coefficient value greater than or equal to 0.7 (sensu Babbitt et al. 2003; Garden et al. 2007) were identified as being highly correlated. No variables were highly correlated, and thus all variables were used in the analyses (Table 2.1).

Models focusing on the influence of environmental variables on abundance and occupancy of L. olongburensis and C. tinnula were constructed a priori. Generalized linear models were used to assess the importance of environmental variables on relative abundance models (using a Poisson link function) and occupancy models (using a binomial link function). I used a generalized ‘rule of thumb’ of n/3 (where n = number of waterbodies sampled) to obtain the maximum number of predictor variables to use in each model (Crawley 2007). Predictor variables included in L. olongburensis and C. tinnula abundance and occupancy models included vegetation cover, predatory fish abundance, pH, salinity, turbidity and depth.

Predation levels can be influenced by the availability of aquatic refuge (i.e. vegetation cover) (Babbitt & Tanner 1998; Kopp et al. 2006).Additionally, it has been proposed that water clarity may impact on predation success (Spieler 2003). Therefore, interaction models between these three variables were included in analysis. Tadpoles of other species were excluded from the models due to the low number of waterbodies in which they were detected. Factors that relate to physiological tolerances often produce a unimodal distribution (e.g. Austin 1999). To detect any unimodal responses, some models included quadratic terms for pH and depth.

Table 2.1: Comparison of habitat characteristics for surveyed waterbodies in wallum habitats of eastern Australia. Spearman correlation coefficients (SCC) were compared for 37 waterbody transects. None of the variables were considered highly correlated (SCC ≥ 0.7). Predatory pH Salinity Depth Turbidity Fish pH 0.290 - - - - Salinity -0.331 -0.281 - - - Depth 0.576 0.178 0.016 - - Turbidity -0.168 -0.357 0.426 -0.042 - Vegetation 0.025 -0.108 -0.110 0.206 -0.086 Cover

Generalised linear models using Akaike’s Information Criterion (AICc to adjust for small sample size (n =37)) were performed in the freeware statistical package R (R Core Development Team , 2011) to determine model ranking and selection. The ‘best’ model was the model with the lowest AICc value (Burnham & Anderson 2002). To determine the ranking of the models, Δi values were calculated, where higher Δi values indicated less accurate models for the given data (Burnham & Anderson 2002; Johnson & Omland 2004).

If a model had a Δi ≤ 2, then there was considerable evidence that the model could be the

“best” model, given the data (Johnson & Omland 2004). If a model had a Δi 2-4 then there was considered to be moderate evidence that the model could be the “best” model, given the data. Akaike Weights (wi) enables greater interpretation of the relative likelihood of a model given the data (Burnham & Anderson 2002; Johnson & Omland 2004). Therefore, each model was assigned a wi, which was used to determine the “probability that model i is the best model for the observed data, given the candidate set of models” (Johnson &

Omland 2004). The closer the wi was to 1, the closer the model for the given data (Burnham & Anderson 2002). To determine the relative importance of variables within models where

Δi < 4 the wi values were summed from all models where the variable of interest occurred (Grueber et al. 2011). The closer the variable of interest was to 1 the higher the importance of the variable. Twenty models were chosen to model L. olongburensis and C. tinnula tadpole relative abundance and occupancy.

Quantile regressions (Cade & Noon 2003; Lancaster & Belyea 2006) were performed in the statistical program R (R Core Development Team, 2011) using the quantreg package (version 4.62) in an attempt to determine the optimal range for factors having a unimodal distribution that were related to L. olongburensis tadpole relative abundance. The number of waterbodies sampled limited the upper and lower quantiles that could be fitted to the data. Therefore, the 0.85 quantile was used to predict the upper limits for relative abundance with 95% confidence intervals. The 0.65 quantile was also plotted to determine shape consistency with the 0.85 quantile. A Gaussian bell-shaped response was assumed for each model.

2.4 Results

Waterbody Characteristics and Rainfall Conditions A total of 37 waterbody transects were surveyed. Five tadpole species were encountered throughout the survey: L. olongburensis, C. tinnula, Litoria gracilenta, Litoria fallax and Litoria cooloolensis. Waterbody transects surveyed had mean pH ranging between 3.00 – 5.04, salinity ranging between 3.32-108.3 parts per million (ppm), turbidity ranging between 0 – 40 nephelometric turbidity units (NTU), mean water depth ranging between 6.2 – 36.2 cm and relative fish abundance between 0 – 67. Tadpoles of L. olongburensis were recorded from 11 waterbodies where the maximum number of individual L. olongburensis tadpoles caught for an individual waterbody transect was nine. Waterbodies containing L. olongburensis tadpoles had a mean pH ranging from 3.40 – 4.34, salinity between 35.82 - 93.72 ppm, turbidity between 0 – 20.4 NTU, mean water depth ranging between 10.2 – 30 cm and relative fish abundance between 0 – 4 individuals. Tadpoles of C. tinnula were recorded from 14 waterbody transects where the maximum number of individual C. tinnula tadpoles caught for an individual waterbody transect was 44. Waterbodies with C. tinnula tadpoles had mean pH ranging between 3.35 – 4.84, salinity between 35.82 – 99.10 ppm, turbidity between 0 - 26.2 NTU, mean water depth ranging between 7.2 – 27.4 cm and relative fish abundance between 0 – 4 individuals. Gambusia holbrooki occurred at four waterbody transects. Tadpoles of L. olongburensis and C. tinnula were not recorded in waterbodies with G. holbrooki. Both L. fallax and L. 49

cooloolensis were recorded from two separate waterbodies. Litoria gracilenta occurred at one waterbody. Tadpoles of L. fallax were not recorded with L. cooloolensis, L. olongburensis or C. tinnula tadpoles. Tadpoles of L. cooloolensis and C. tinnula were found co-occurring in 3 waterbodies.

The average rainfall for the months between December 2009 – March 2010 differed between weather stations, with above average rainfall being recorded for the Beerburrum Forest Reserve (+219mm) and Byron Bay Cap Byron Lighthouse (+202.2mm) weather stations and below average rainfall for the Double Island Point Lighthouse (-210.9mm) and Wooli Beach (-87.9mm) weather stations.

Abundance

Three models for L. olongburensis relative abundance had a Δi < 2. The weighting of the best models for L. olongburensis tadpole relative abundance was 99.8%, suggesting that the other models compared poorly (Table 2.2). The best model contained pH, depth and turbidity, with all variables having strong relative importance (Table 2.3). Turbidity was the only factor within the best model that was negatively associated with the relative abundance of L. olongburensis tadpoles (Table 2.2). The other two models contained all variables measured. Other variables within these models had lower relative variable importance when compared with pH, depth and turbidity (Table 2.3).

Unlike L. olongburensis, C. tinnula tadpole relative abundance was best explained (weighting of 100%), by the model with all predictor factors (Table 2.2).

Occupancy

Two models for L. olongburensis tadpole occupancy had a Δi < 2 while only one model had a Δi between 2-4. The combined weighting of the models with a Δi < 2 for L. olongburensis occupancy was 70%, suggesting that the other models compared poorly given the data.

Table 2.2: Comparison of waterbody characteristics associated with the relative abundance and occupancy of L. olongburensis or C. tinnula tadpoles in eastern Australia. Aikiki models with Δi values less than 4 are presented. + indicates a positive relationship while – indicates a negative relationship to L. olongburensis or C. tinnula tadpole relative abundance or occupancy. Variables with a 2 indicate a unimodal distribution with L. olongburensis or C. tinnula tadpole relative abundance or occupancy.

Model AICc Δi wi Litoria olongburensis tadpole relative abundance pH – pH2 + Depth – Depth2 – Turbidity 126.2 0 0.444 pH - pH2 + Salinity + Depth - Depth2 - 126.9 0.64 0.322 Turbidity - % Cover pH - pH2 + Salinity + Depth - Depth2 - 127.5 1.30 0.232 Turbidity - Predatory Fish - % Cover Litoria olongburensis tadpole occupancy pH – pH2 43.2 0 0.463 pH – pH2 + Depth – Depth2 44.5 1.34 0.237 Depth – Depth2 46.9 3.69 0.073 Crinia tinnula tadpole relative abundance pH - pH2 - Salinity + Depth - Depth2 + 472.4 0 1 Turbidity - Predatory Fish - % Cover Crinia tinnula tadpole occupancy (-) Predatory Fish 48.1 0 0.257 (-) Depth 48.8 0.67 0.184 (-) Predatory Fish + Turbidity 49.9 1.78 0.106 Depth – Depth2 50.2 2.10 0.09 Turbidity 51.7 3.52 0.044 Predatory Fish * Depth 51.8 3.66 0.041 Salinity 51.9 3.77 0.039 Predatory Fish * Turbidity 51.9 3.8 0.038 (-) % Cover 52 3.84 0.038 (-) % Cover – Predatory Fish + Turbidity 52.1 3.99 0.035

Variable Estimate S.E. † Confidence Rel. var. Interval imp.‡ Litoria olongburensis abundance pH 59.5 16.5 26.02, 92.994 1 pH2 -7.47 2.08 -11.7, -3.247 1 Depth 0.4 0.127 -0.143, 0.663 1 Depth2 -0.01 -0.003 -0.018, -0.004 1 Turbidity -0.026 -0.03 -0.087, 0.034 1 % Cover -0.017 0.009 -0.036, 0.002 0.55 Salinity 0.0243 0.012 -0.0002, 0.049 0.55 Predatory Fish -0.137 -0.096 -0.334, 0.06 0.23 Litoria olongburensis occupancy pH 54.689 34.155 -14.24, 123.62 0.76 pH2 -7.33 4.158 -15.76, 1.1 0.72 Depth 0.625 0.393 -0.167, 1.418 0.36 Depth2 -0.018 0.009 -0.037, -0.004 0.33 Crinia tinnula occupancy Predatory Fish -0.395 0.525 -1.459, 0.67 0.48 Depth -0.013 0.166 -0.346, 0.319 0.38 Depth2 -0.006 0.007 -0.02, 0.007 0.11 Turbidity 0.031 0.038 -0.046, 0.108 0.28 Salinity 0.01 0.016 -0.023, 0.043 0.07 % Cover -0.008 0.015 -0.038, 0.022 0.07 Predatory Fish * Turbidity -0.027 0.054 -0.137, 0.082 0.04 Predatory Fish * Depth 0.022 0.041 -0.061, 0.104 0.04 † Standard Error ‡ Relative variable importance

Models with a Δi < 2 contained pH and depth, with pH having the highest relative variable importance (Table 2.3). Both variables were unimodel in their influence on L. olongburensis occupancy (Table 2.2).

Three models for C. tinnula tadpole occupancy had a Δi < 2 while seven models had a

Δi between 2-4. The combined weighting for models with a Δi < 2 for C. tinnula tadpole occupancy was 54.7%. These best models contained predatory fish, turbidity and depth. Predatory fish and depth had the highest relative variable importance when compared with other variables (Table 2.3). Predatory fish and depth were negatively associated with C. tinnula tadpole occupancy, while turbidity was positively associated with C. tinnula tadpole occupancy (Table 2.2).

Optimal Range

Determining an optimal range for factors influencing L. olongburensis tadpole abundance is difficult due to large 95% confidence intervals in the maximum (0.85 quantile) abundance models. However, quantile regressions showed a unimodal response between L. olongburensis tadpole relative abundance and water depth and pH (Figure 2.2 and 2.3). Confidence intervals for pH are large for the 0.85 quantile and are lowest towards the lower and upper pH ranges. Additionally, confidence intervals are smallest towards the deeper water depths for the 0.85 quantile and largest towards shallower water depths.

The Gaussian distribution fits well for the 0.85 and 0.65 pH quantiles, with p < 0.001 for all co-efficient values. The intercept of the 0.85 and the 0.65 water depth quantiles have p > 0.05, indicating that intercepts for these quantiles do not fit a Gaussian distribution. However the Depth and Depth2 for the 0.85 quantiles fit a Gaussian distribution with co-efficient values having p < 0.05. The 0.65 water depth quantile lacks a Gaussian distribution fit, with p > 0.05 for all co-efficient values (Table 2.4).

Figure 2.2: ‘Jitter’ plots for quantile regressions of the 0.85(a) and 0.65(b) quantiles (solid line) and the 95% confidence intervals (dotted lines) for mean waterbody pH and relative abundance of Litoria olongburensis tadpoles. Circles represent waterbody transects.

Quantile values for Litoria olongburensis and pH Parameter Estimate S.E.† t p-value 0.85 quantile Intercept -29.74462 7.03260 -4.22953 0.00017 pH 15.81659 3.40438 4.64595 0.00005 pH2 -1.96724 0.40236 -4.88921 0.00002 0.65 quantile Intercept -29.28921 12.66667 -2.31231 0.02695 pH 15.09734 6.11008 2.47089 0.01865 pH2 -1.84790 0.71455 -2.58611 0.01416 Quantile values for Litoria olongburensis and water depth

Parameter Estimate SE† t p-value

0.85 quantile Intercept -1.66692 1.05192 -1.58465 0.12230 Depth 0.35474 0.08616 4.11721 0.00023 Depth2 -0.00853 0.00167 -5.11025 0.00001 0.65 quantile Intercept -0.79584 0.50358 -1.58036 0.12328 Depth 0.15177 0.08050 1.88529 0.06796 Depth2 -0.00378 0.00186 -2.02956 0.05029 † Standard Error

2.5 Discussion

I present the first quantitative assessment of environmental variables associated with the relative abundance and occupancy of tadpoles within waterbodies of wallum heathland of eastern Australia. As expected, waterbodies in wallum vegetation were dominated by tadpoles of the “acid frog” (L. olongburensis and C. tinnula), with the abundance and occupancy of these species associated with either water chemistry variables or predatory fish. Additionally, tadpoles of L. fallax and L. cooloolensis were also recorded from two independent waterbodies. The low occupancy recorded for L. cooloolensis is likely due to surveys only being conducted within a proportion of L. cooloolensis distributional range. However, the low occupancy of L. fallax tadpoles is likely explained by this species inability to successfully metamorphose in acidic waters (pH = 3.5) (Meyer 2004).

Variables influencing abundance and occupancy

Waterbody pH and water depth were associated with the relative abundance and occupancy of L. olongburensis tadpoles. The range of pH values of waterbodies where L. olongburensis tadpoles were found falls within the pH ranges of other studies that found L. olongburensis adults between pH 3.8 – 4.6 (Hopkins unpublished data), 3.5 – 5.2 (Hero unpublished data) and 3.11 – 5.02 (Shuker unpublished data). While L. olongburensis tadpole pH tolerance has never been tested, two other ‘acid’ frogs (C. tinnula and L. cooloolensis) have successfully metamorphosed when exposed to pH waters of 3.5, 4.5 and 6.5 (Meyer 2004). It is expected that L. olongburensis tadpoles would also be able to metamorphose when exposed to these pH levels as tadpoles of L. olongburensis co-occurred with C. tinnula and L. cooloolensis tadpoles within this study.

Large confidence intervals and the quantile ranges extending over a majority of the response variables tested (pH and depth) give a guide to the optimal range for L. olongburensis tadpole abundance. Despite large confidence intervals, the 0.85 and 0.65 quantile for pH showed a unimodal response, suggesting that pH is influencing L. olongburensis as a quadratic, nonlinear response. The unimodal relationship observed

suggests that competition, predation or a physiological intolerance is limiting this species at the upper or lower bounds of the unimodal relationship. Acidity intolerance has been recorded for numerous amphibian species, where death occurs to tadpoles outside of their pH tolerance range (reviews in Pierce 1985 & Freda 1986; Meyer 2004). Therefore, physiological intolerance of L. olongburensis tadpoles to pH levels occurring at the lower limits of the pH range within this study is probable. However, experimental tests on L. olongburensis tadpole tolerance using pH levels lower than those used by Meyer (2004) are required to confirm this conclusion. Towards the upper limits of the pH quantile regressions it is possible that competition from ‘non-acid’ frog species (i.e. L. fallax) may influence the abundance of L. olongburensis tadpoles. This is difficult to confirm in this study as only two waterbodies recorded a potential competitor (L. fallax) and pH within waterbodies surveyed did not exceed 5.21. It is unlikely that ‘non-acid’ frog species would successfully compete and exclude or reduce L. olongburensis tadpoles in the lower pH ranges as ‘non-acid’ frog tadpoles (i.e. L. fallax and Crinia parinsignifera) have 100% hatchling mortality or 0% hatching success when exposed to pH levels of 3.5 (Meyer 2004).

The 0.85 quantile for water depth showed a unimodal response, suggesting that water depth is influencing L. olongburensis in a quadratic and not a linear response. Quantile regression models for L. olongburensis tadpoles and water depth showed that tadpole relative abundance will increase with water depth until water depth reaches approximately 22cm, after which tadpole relative abundance will decrease. Water depth was used as an indicator of hydroperiod, with deeper waterbodies representing waterbodies with longer hydroperiods. The reduced number of L. olongburensis tadpoles in deeper waterbodies is likely associated with an increase predatory fish, which need a more permanent hydroperiod for population persistence, or a change in the predator assemblage. Predatory fish were not in the best models explaining L. olongburensis tadpole abundance or occupancy. However, predatory fish had the highest correlation coefficient with depth when compared with other measured variables. Additionally, predatory fish only co-occurred with L. olongburensis tadpoles in two waterbodies, suggesting either coexistence between predatory fish and L. olongburensis tadpoles is low or abundance of L. olongburensis tadpoles is lower when coexisting with predatory fish. Aquatic fauna assemblages (Welborn et al. 1996;

Babbitt et al. 2003; Jocqué et al. 2007; Richter-Boix et al. 2007; Fernandes et al. 2010) and an increase in predator size (Welborn et al. 1996; Richter-Boix et al. 2007) are associated with increased hydroperiod or water depth. Therefore, an increase in hydroperiod/depth may result in the number of L. olongburensis tadpoles being reduced due to inadequate anti-predator strategies to larger predators (i.e. tadpole size may be an inadequate anti-predator strategy to predators with large mouth-gapes) occurring in deeper sections of a waterbody. Alternatively, deeper waterbodies may have reduced the probability of detecting tadpoles due to an increase in the volume of water. Decreasing hydroperiod has been shown to induce early metamorphosis (Loman 1999; Lane & Mahony 2002; Loman 2002; Márquez-García et al. 2010) and tadpoles of L. olongburensis may have metamorphosed in shallower waterbodies in an attempt to metamorphose before waterbody desiccation. Alternatively, shallower waterbodies may never have contained tadpoles of L. olongburensis. Water height has been found to influence oviposition-site selection (Goldberg et al. 2006) and adult L. olongburensis may have chosen deeper waterbodies with a lower desiccation risk to lay their eggs.

Turbidity was positively associated with C. tinnula tadpole occupancy and negatively associated with L. olongburensis tadpole abundance. Numerous factors may have caused turbidity, including natural organic acids (NOA), which produce dark, tannin- stained waters (Barth & Wilson 2010), and/or suspended sediment (Walling 1977). NOA has been attributed to an increase in growth and locomotor performance in Limnodynastes peronii tadpoles (Barth & Wilson 2010) and other protective mechanisms in various aquatic species (Gonzalez et al. 2002; Wood et al. 2003) when exposed to acidic conditions. Similar, positive, processes could be occurring with C. tinnula tadpoles. Alternatively, reduced tadpole numbers have been observed or proposed with an increased turbidity in previous surveys (Gillespie 2002; Knutson et al. 2004; Schmutzer et al. 2008), with several hypothesis suggested as to the cause of the decline, including suffocation of tadpole eggs (hypothesised from suffocation of fish eggs in waters with high turbidity) or reduced ability to forage (Schmutzer et al. 2008). Similar, negative, processes could be occurring with L. olongburensis tadpoles.

There were several waterbody transects that failed to detect L. olongburensis tadpoles despite being within the optimal pH and water depth ranges. The adult surveys that

coincided with these surveys found a majority of waterbodies to have L. olongburensis adults present (Shuker unpublished data). The absence of tadpoles from these waterbodies suggests that eggs may have been deposited but did not survive long enough, or were in such low densities that they were not detected. As mentioned previously, females also have the ability to choose their oviposition site (Goldberg et al. 2006) and females may have chosen alternative spawning sites due to an unsuitable environmental factor. Alternatively, tadpoles occurring in larger, deeper waterbodies may have been missed due to an increase in the sampling area in larger waterbodies and tadpoles potentially being dispersed over a larger area.

Predatory fish were low in occupancy and abundance throughout the survey. Hence, the following conclusions must be interpreted with some caution. Having said this, the given data showed that predatory fish were negatively associated with C. tinnula tadpole occupancy. Predatory fish have the ability to exclude tadpole species that lack adequate anti-predator defences from waterbodies (Kats et al. 1988; Hero et al. 1998, 2001; Gillespie 2001). Species that lack efficient anti-predatory fish defences (e.g. unpalatability (Kats et al. 1988; Hero et al. 2001) or reduced mobility (Woodward 1983) may therefore be confined to waterbodies with ephemeral hydroperiods where fish are often excluded due to waterbody desiccation (Heyer et al. 1975). Therefore, the negative association between predatory fish and C. tinnula tadpole occupancy is likely explained by lack of efficient anti-predatory fish strategies.

The model that best explained C. tinnula relative abundance was the model with all variables, suggesting there are complex interactions occurring between all variables and C. tinnula tadpole relative abundance. Further survey work and experimentation are required to determine the key factors influencing C. tinnula tadpole relative abundance.

Implications for Conservation

The lack of co-occurrence of C. tinnula tadpoles and low co-occurrence of L. olongburensis with fish suggests that introduction of fish (both native and exotic) into wallum waterbodies would be detrimental to populations of these two threatened ‘acid’ frog species. Therefore, it is imperative that introduction of fish into wallum

waterbodies is avoided. For example, establishment of permanent waterbodies, where fish would likely have the highest chance of population persistence, may result in fish populations spreading to fish naive waterbodies during flooding events. This scenario has to be taken into consideration if construction of ‘permanent’ waterbodies are undertaken to reduce the effects of climate change (Shoo et al. 2011) on ‘acid’ frog populations. Additionally, establishment of permanent waterbodies is of particular concern if waterbodies contain G. holbrooki, a known introduced, predatory fish within Australia (Pyke & White 2000; Reynolds 2009; Hunter et al. 2011), which is capable of spreading between waterbodies in surface water runoff (Baber et al. 2002). The impact of G. holbrooki on L. olongburensis and C. tinnula abundance and presence is difficult to assess from this study, as only four waterbodies contained G. holbrooki. However, G. holbrooki did not occur in the same waterbodies with L. olongburensis and C. tinnula tadpoles, suggesting that G. holbrooki could negatively influence these amphibian populations.

The low occupancy of L. fallax tadpoles across waterbodies indicates L. fallax do not breed successfully in most waterbodies within wallum heathland. This is of particular importance for conservation of waterbodies within wallum heathland, as changed waterbody chemistry could result in conditions that are favourable for the generalist L. fallax, which has been suggested as a competitive species to the threatened specialist L. olongburensis (Meyer et al. 2006). Stormwater and runoff from urban areas and golf courses have been proposed as potentially altering wallum waterbody chemistry (Meyer et al. 2006) and, therefore, these anthropogenic disturbances need to be monitored and diverted away from wallum waterbodies to maintain natural waterbody chemistry.

Waterbody transect length gave an indication of waterbody size, with the smallest waterbody surveyed being no less than 48 meters from one waterbody edge to the other. Waterbodies may fluctuate in size during the year and results from the study can only be attributed to scenarios that were occurring within the survey period. However, smaller waterbodies then those detected in this study outside of protected areas or within natural habitats at different times of year may also be suitable for L. olongburensis and C. tinnula. Furthermore, additional surveys are required to detect winter breeding species, and determine how abundances and occupancy of the species outside of protected areas.

Despite these limitations, these results demonstrate a strong association with low pH and the threatened amphibian species recorded. Additional environmental variables associated with the distribution and abundance of amphibian tadpole assemblages within oligotrophic, acidic waterbodies are complex and varies between species. Therefore, conservation requirements of wallum heathland amphibian species need to be assessed separately for each species, and encompass a range of waterbodies where environmental variables are within the species optimal tolerance range. Furthermore, pH and water depth influencing relative abundance and occupancy of L. olongburensis tadpoles indicate that either measurement can be used when assessing waterbody suitability for tadpoles of the specialist species, L. olongburensis.

Acknowledgements

I would like to thank the Griffith School of Environment and FKP Ltd for providing project funding for this chapter. I also thank Sonya Clegg, Katrin Lowe, Donna Treby, two anonymous reviewers and the associate editor of Austral Ecology for feedback on this chapter (sections of which have been accepted for publication in Austral Ecology*). The submitted manuscript of this chapter was co-authored by Jon Shuker, Gregory Lollback, Guy Castley and Jean-Marc Hero. Griffith Animal Ethics number ENV/17/08, New South Wales Department of Environment and Climate Change Scientific License number S21717 and Queensland Department of Environment and Resource Management permit number WITK05620308 were obtained for this work conducted in this chapter.

*Simpkins, C.A., Shuker, J.D., Lollback, G.W., Castley, J.G. and Hero, J.-M. (2013). Environmental variables associated with the distribution and occupancy of habitat speclialist tadpoles in naturally acidic, oligotrophic waterbodies. Austral Ecology, In press.

2.6 References

Austin M.P. (1999) The potential contribution of vegetation ecology to biodiversity research. Ecography 22, 465-484.

Babbitt K.J. & Tanner G.W. (1998) Effects of cover and predator size on survival and development of Rana utricularia tadpoles. Oecologia, 114.

Babbitt K.J., Baber M.J. & Tarr T.L. (2003) Patterns of larval amphibian distribution along a wetland hydroperiod gradient. Can. J. Zool. 81, 1539-1552.

Baber M.J., Childers D.L., Babbitt K.J. & Anderson D.H. (2002) Controls on fish distribution and abundance in temporary wetlands. Can. J. Fish. Aquat. Sci. 59, 1441-1450.

Baber M.J., Fleishman E., Babbitt K.J. & Tarr T.L. (2004) The relationship between wetland hydroperiod and nestedness patterns in assemblages of larval amphibians and predatory macroinvertebrates. Oikos 107, 16-27.

Barth B.J. & Wilson R.S. (2010) Life in acid: interactive effects of pH and natural organic acids on growth, development and locomotor performance of larval striped marsh frogs (Limnodynastes peronii). J. Exp. Biol. 213, 1293-1300.

Berkelmans R. & Willis B.L. (1999) Seasonal and local spatial patterns in the upper thermal limits of corals on the inshore Central Great Barrier Reef. Coral Reefs 18, 219-228.

Burnham K.P. & Anderson D.R. (2002) Model Selection and Multimodel Inference: A Practical Information-Theoretic Approach. Springer, New York.

Cade B.S. & Noon B.R. (2003) A gentle introduction to quantile regression for ecologists. Front. Ecol. Environ. 1, 412-420.

Chinathamby K., Reina R.D., Bailey P.C.E. & Lees B.K. (2006) Effects of salinity on the survival, growth and development of tadpoles of the brown tree frog, Litoria ewingii. Aust. J. Zool. 54, 97-105.

Christy M.T. & Dickman C.R. (2002) Effects of salinity on tadpoles of the green and golden bell frog (Litoria aurea). Amphibia-Reptilia 23, 1-11.

Crawley M.J. (2007) The R Book. Wiley, England.

Fernandes I.M., Machado F.A. & Penha J. (2010) Spatial pattern of a fish assemblage in a seasonal tropical wetland: effects of habitat, herbaceous plant biomass, water depth, and distance from species sources. Neotrop. Ichthyol. 8, 289-298.

Freda J. (1986) The influence of acidic pond water on amphibians: A review . Water Air Soil Poll. 30, 439-450.

Freda J. & Taylor D.H. (1992) Behavioral response of amphibian larvae to acidic water. J. Herpetol. 26, 429-433.

Garden J.G., McAlpine C.A., Possingham H.P. & Jones D.N. (2007) Habitat structure is more important than vegetation composition for local-level management of native terrestrial reptile and small mammal species living in urban remnants: A case study from Brisbane, Australia. Austral Ecol. 32, 669-685.

Gillespie G.R. (2001) The role of introduced trout in the decline of the spotted tree frog (Litoria spenceri) in south-eastern Australia. Biol. Conserv. 100, 187-198.

Gillespie, G.R. (2002). Impacts of sediment loads, tadpole density, and food type on the growth and development of tadpoles of the spotted tree frog Litoria spenceri: an in-stream experiment. Biol. Conser. 106, 141-150.

Goldberg F.J., Quinzio S. & Vaira M. (2006) Oviposition-site selection by the toad Melanophryniscus rubriventris in an unpredictable environment in Argentina. Can. J. Zool. 84, 699-705.

Gonzalez R.J., Wilson R.W., Wood C.M., Patrick M.L. & Val A.L. (2002) Diverse strategies for ion regulation in fish collected from the ion-poor, acidic Rio Negro. Physiol. Biochem. Zool. 75, 37-47.

Gosner K.L. & Black I.H. (1957) The effects of acidity on the development and hatching of New Jersey Frogs. Ecology 38, 256-262.

Grueber C.E., Nakagawa S., Laws R.J. & Jamieson I.G. (2011) Multimodel inference in ecology and evolution: challenges and solutions. J. Evolution. Biol. 24, 699-711.

Hecnar S.J. & M’Closkey R.T. (1997) The effects of predatory fish on amphibian species richness and distribution. Biol. Conserv. 79, 123-131.

Hero J.-M., Gascon C. & Magnusson W.E. (1998) Direct and indirect effects of predation on tadpole community structure in the Amazon rainforest. Aust. J. Ecol. 23, 474-482.

Hero J.-M., Magnusson W.E., Rocha C.F.D. & Catterall C.P. (2001) Antipredator defenses influence the distribution of amphibian prey species in the Central Amazon rain forest. Biotropica 33, 131-141.

Heyer W.R. (1973) Ecological interactions of frog larvae at a seasonal tropical location in Thailand. J. Herpetol. 7, 337-361.

Heyer W.R., McDiarmid R.W. & Weigmann D.L. (1975) Tadpoles, predation and pond habitats in the tropics. Biotropica 7, 100-111.

Hossie T.J. & Murray D.L. (2010) You can’t run but you can hide: refuge use in frog tadpoles elicits density-dependent predation by dragonfly larvae. Oecologia 163, 395-404. 64

Hunter D.A., Smith M.J., Scroggie M.P. & Gilligan D. (2011) Experimental examination of the potential for three introduced fish species to prey on tadpoles of the endangered Booroolong Frog, Litoria booroolongensis. J. Herpetol. 45, 181- 185.

Ingram G.J. & Corben C.J. (1975) The frog fauna of North Stradbroke Island, with comments on the “acid” frogs of the wallum. Proc. R. Soc. Queensl. 86, 49-54.

IUCN 2011. (2011) IUCN Red List of threatened species. Available from http://www.iucnredlist.org (accessed July 2011).

Jocqué M., Graham T. & Brendonck L. (2007) Local structuring factors of invertebrate communities in ephemeral freshwater rock pools and the inﬂuence of more permanent water bodies in the region. Hydrobiologia 592, 271-280.

Johnson J.B. & Omland K.S. (2004) Model selection in ecology and evolution. Trends Ecol. Evol. 19, 101-8.

Kats L.B., Petranka J.W. & Sih A. (1988) Antipredator defenses and the persistence of amphibian larvae with fishes. Ecology 69, 1865-1870.

Knutson M.G., Richardson W.B., Reineke D.M., Gray B.R., Parmelee J.R. & Weick S.E. (2004) Agricultural ponds support amphibian populations. Ecol. Appl. 14, 669-684.

Kopp K., Wachlevski M. & Eterovick P.C. (2006) Environmental complexity reduces tadpole predation by water bugs. Can. J. Zool. 84, 136-140.

Krebs C.J. (2009) Ecology: Sixth Edition. Benjamin Cummings, United States of America.

Lancaster J. & Belyea L.R. (2006) Defining the limits to local density: alternative views of abundance-environment relationships. Freshwater Biol. 51, 783-796.

Lane M.J. & Mahony S.J. (2002) Larval anurans with synchronous and asynchronous development periods: contrasting responses to water reduction and predator presence. J. Anim. Ecol. 71, 780-792.

Lewis B.D. & Goldingay R.L. (2005) Population monitoring of the vulnerable wallum sedge frog (Litoria olongburensis) in north-eastern New South Wales. Aust. J. Zool. 53, 185-194.

Loman J. (1999) Early metamorphosis in common frog Rana temporaria tadpoles at risk of drying: an experimental demonstration. Amphibia-Reptilia 20, 421-430.

Loman J. (2002) Temperature, genetic and hydroperiod effects on metamorphosis of brown frogs Rana arvalis and R. temporaria in the field. J. Zoo. 258, 115-129.

Meyer E. (2004) Acid adaptation and mechanisms for softwater acid tolerance in larvae of anuran species native to the “Wallum” of east Australia. PhD Thesis, University of Queensland.

Meyer E., Hero J-M., Shoo L. & Lewis B. (2006) National recovery plan for the wallum sedgefrog and other wallum-dependent frog species. Report to Department of the Environment and Water Resources, Canberra. Queensland Parks and Wildlife Service, Brisbane.

Mokany A. & Shine R. (2002) Competition between tadpoles and mosquito larvae. Oecologia 135, 615-620.

Moreira L.F.B., Machado I.F., Garcia T.V. & Maltchik L. (2010) Factors influencing anuran distribution in coastal dune wetlands in southern Brazil. J. Nat. Hist. 44, 1493-1507.

Morin P.J. (1986) Interactions between intraspecific competition and predation in an amphibian predator-prey system. Ecology 67, 713-720.

Márquez-García M., Correa-Solís M. & Méndez M.A. (2010) Life-history trait variation in tadpoles of the warty toad in response to pond drying. J. Zoo. 281, 105-111.

Persson M., Räsänen K., Laurila A. & Merilä J. (2007) Maternally determined adaptation to acidity in Rana arvalis: Are laboratory and field estimates of embryonic stress tolerance congruent? Can. J. Zool. 85, 832-838.

Pierce B.A. (1985) Acid tolerance in amphibians. BioScience 35, 239-243.

Pyke A.W. & White G.H. (2000) Factors influencing predation on eggs and tadpoles of the endangered Green and Golden Bell Frog Litoria aurea by the introduced Plague Minnow Gambusia holbrooki. Aust. Zoo. 31, 496-505.

Reynolds S.J. (2009) Impact of the introduced Poeciliid Gambusia holbrooki on amphibians in southwestern Australia. Copeia 2009, 296-302.

Richter-Boix A., Llorente G.A. & Montori A. (2007) A comparative study of predator- induced phenotype in tadpoles across a pond permanency gradient. Hydrobiologia 583, 43-56.

Rios-López N. (2008) Effects of increased salinity on tadpoles of two anurans from a Caribbean coastal wetland in relation to their natural abundance. Amphibia-Reptilia 29, 7-18.

Sadinski W.A. & Dunson W.J. (1992) A multilevel study of effects of low pH on amphibians of temporary ponds. J. Herpetol. 26, 413-422.

Saidapur S.K., Veeranagoudar D.K., Hiragond N.C. & Shanbhag B.A. (2009) Mechanism of predator–prey detection and behavioral responses in some anuran tadpoles. Chemoecology 19. 66

Schmutzer A.C., Gray M.J., Burton E.C. & Miller D.L. (2008) Impacts of cattle on amphibian larvae and the aquatic environment. Freshwater Biol. 53, 2613-2625.

Schofield P.J. & Nico L.G. (2009) Salinity tolerance of non-native Asian swamp eels (Teleostei: Synbranchidae) in Florida, USA: comparison of three populations and implications for dispersal. Environ. Biol. Fish. 85, 51-59.

Shuker, J.D., Hero, J.-M., 2012. Perch substrate use by the threatened wallum sedge frog (Litoria olongburensis) in wetland habitats of mainland eastern Australia. Aust. J. Zoo. http://dx.doi.org/10.1071/ZO12030.

Shoo L.P., Olson D.H., McMenamin S.K., Murray K.A., Van Sluys M., Donnelly M.A., et al. (2011) Engineering a future for amphibians under climate change. J. Appl. Ecol. 48, 487-492.

Smith M.J., Schreiber E.S.G., Scroggie M.P., Kohout M., Ough K., Potts J., et al. (2007) Associations between anuran tadpoles and salinity in a landscape mosaic of wetlands impacted by secondary salinisation. Freshwater Biol. 52, 75-84.

Snodgrass J.W., Bryan Jr. A.L. & Burger J. (2000) Development of expectations of larval amphibian assemblage structure in southeastern depression wetlands. Ecol. Appl. 10, 1219-1229.

Spieler, M. (2003). Risk of predation affects aggregation size: a study with tadpoles of Phrynomantis microps (Anura: Microhylidae). Anim. Behav.65, 179-184.

Twomey E., Morales V. & Summers K. (2008) Evaluating condition-specific and asymmetric competition in a species-distribution context. Oikos 117, 1175-1184.

Vonesh J.R., Kraus J.M., Rosenberg J.S. & Chase J.M. (2009) Predator effects on aquatic community assembly: disentangling the roles of habitat selection and post- colonization processes. Oikos 118, 1219-1229.

Walling D.E. (1977) Assessing the accuracy of suspended sediment rating curves for a small basin. Water Resour. Res. 13, 531-538.

Welborn G.A., Skelly D.K. & Werner E.E. (1996) Mechanisms creating community structure across a freshwater habitat gradient. Annu. Rev. Ecol. Syst. 27, 337-363.

Wells K.D. (2007) The Ecology and Behavior of Amphibians. The University of Chicago Press, United States of America.

Wilbur H.M. (1987) Regulation of structure in complex systems: experimental temporary pond communities. Ecology 68, 1437-1452.

Wiltshire D.J. & Bull C.M. (1977) Potential competitive interactions between larvae of Pseudophryne bibroni and P. Semimarmorata (Anura: Leptodactylidae). Aust. J.l of Zool. 25, 449-454.

Wood C.M., Matsuo A.Y.O., Wilson R.W., Gonzalez R.J., Patrick M.L., Playle R.C., et al. (2003) Protection by natural blackwater against disturbances in ion fluxes caused by low pH exposure in freshwater stingrays endemic to the Rio Negro. Physiol. Biochem. Zool.76, 12-27.

Woodward B.D. (1983) Predator-prey interactions and breeding-pond use of temporary- pond species in a desert anuran community. Ecology 64, 1549-1555.

Chapter 3 - Suitability of anthropogenic waterbodies for amphibians associated with naturally acidic, oligotrophic environments and environmental variables influencing their distribution

3.1 Abstract

Habitat destruction is a key threatening process to amphibians outside of protected areas. Anthropogenically modified or waterbodies can be used to compensate for habitat loss. Several amphibian species utilise artificial waterbodies as compensatory habitats, or opportunistically. Use of anthropogenic/modified waterbodies by adult and tadpoles of amphibian species that are both associated and not associated with naturally acidic, oligotrophic waterbodies was measured. Additionally, the environmental factors influencing the anuran assemblage, with a focus on Litoria olongburensis (a threatened species associated with naturally acidic, oligotrophic waterbodies) and Litoria fallax (a common, potentially competitive species of L. olongburensis) was examined. Nine road trenches/ditches, eight artificial ‘lakes’, six abandoned golf course waterbodies and thirteen natural waterbodies were surveyed for anuran adults and tadpoles during the summer/spring period of 2011/2012. Water chemistry, aquatic predators and vegetation types were also measured. Water chemistry and anuran species richness differed among waterbody types, with two of three threatened anuran species present in both natural and anthropogenic/modified waterbodies. The anuran assemblage was influenced by several environmental variables, including pH, turbidity, salinity and % cover of certain vegetation types. Relative abundance of adults and tadpole occupancy of L. olongburensis was highest in natural waterbodies, while adult relative abundance of L. fallax was highest within artificial lakes. Additionally, L. olongburensis relative abundance was positively associated with increased sedge density and negatively associated with increased water pH. This study demonstrates that differing levels of water chemistry factors and vegetation density influence the amphibian assemblage within these environments. Effective conservation of all anuran species would be enhanced by conserving a variety of waterbody types, however, natural waterbodies provide the best conservation protection for threatened species including L. 69

olongburensis and C. tinnula. Environmental managers creating habitat offsets should therefore critically assess the quality of constructed habitat for specialist anuran species.

3.2 Introduction

Amphibian species are declining globally (Stuart et al., 2004), with habitat loss and modification being one of the main threatening factors (Collins and Storfer, 2003; Beebee and Griffiths, 2005). Despite this, modified or anthropogenic waterbodies can be used to compensate for habitat loss (e.g. Mazerolle et al., 2006; Ruhí et al., 2012), with several amphibian species within Australia (reviewed in Hazell, 2003; Hazell et al., 2004; Lemckert et al., 2006), North America (Monello and Wright, 1999; Brand and Snodgrass, 2010; Brown et al., 2012) and Europe (Rannap et al., 2009; Brown et al., 2012; Ruhí et al., 2012) utilising these habitats. However, species richness and assemblages can differ between these and natural waterbodies (Hazell et al., 2004), potentially due to differences in environmental factors that influence amphibian species distributions within waterbodies.

Several environmental factors can influence the distribution of amphibians within waterbodies, including aquatic predators (Kats et al., 1988; Hecnar and M’Closkey, 1997; Hero et al., 2001; Vonesh et al., 2009) or competitors (Morin, 1986; Wilbur, 1987; Mokany and Shine, 2002; Twomey et al., 2008). Tadpole intolerance to water chemistry variables, including pH (reviewed in Pierce, 1985; Sparling, 2010) and salinity (Strahan, 1957; Christy and Dickman, 2002; Chinathamby et al., 2006; Rios- López, 2008) may also exclude amphibian species from a waterbody due to water chemistry intolerances. Additionally, turbidity has been found to influence amphibian species richness (Hecnar and M’Closkey, 1996) and may influence the structure of other amphibian assemblages. Furthermore, the abundance or emergence of particular vegetation species (Lemckert et al., 2006; Shuker, 2012), or the proportion of the water margin with emergent vegetation (Hazell et al., 2004) Lemckert et al., 2006, may also influence amphibian usage or species abundances within waterbodies. Pond isolation may also influence amphibian species richness within Australian ponds (Smallbone et al., 2011).

Reductions in the availability of coastal waterbodies have occurred globally (e.g. Turner, 1990; Levin et al., 2009), due to human population growth and urban expansion. Within Australia, the majority of freshwater, coastal waterbodies situated along the eastern seaboard between Fraser Island, QLD, and Jervis Bay, NSW, are unique as they are both naturally oligotrophic (Hines et al., 1999) and acidic (Coaldrake, 1961; Griffith et al., 2008, Hines and Meyer, 2011). The ‘wallum’ waterbodies of these coastal areas, characterised by Banksia woodlands, sedgeland, heathland and Melaleuca swamps (Hines et al., 1999; Griffith et al., 2003), also support stable populations of threatened ‘acid’ frog species (Ingram and Corben, 1975; Hines et al., 1999; Lewis and Goldingay, 2005). However, populations occurring outside of protected areas are at risk from habitat loss (Hines et al., 1999), with more than 30% of original heathland and Melaleuca cover in south-east QLD cleared between 1974-1989 (Catterall and Kingston, 1993, cited in Hines et al., 1999). Furthermore, areas within this range have one of the highest human growth rates within Australia (Hines et al., 1999; Garden et al., 2010). Wallum habitats also contain modified and anthropogenic waterbodies (Simpkins pers. obs.) which may be utilised by ‘acid’ frog species and aid in reducing the impacts of habitat loss. Despite the numerous threats to these species and the presence of anthropogenic or modified waterbodies in their habitat, no peer-reviewed studies have been undertaken to determine if anurans normally associated with naturally acidic, oligotrophic waterbodies use and successfully reproduce within anthropogenic or modified waterbodies.

Consequently, this chapter aims to compare use of natural and anthropogenic waterbodies by adult and tadpole anuran species around and within coastal wallum of eastern Australia. Furthermore, the study examines which environmental variables (aquatic predators, water chemistry variables, vegetation type) influence the anuran assemblage and the relative abundance of the threatened wallum sedge frog, Litoria olongbureneis, and a potential competitor, Litoria fallax (Meyer et al., 2006), across the natural-modified landscape. Information obtained from this study will aid environmental managers creating habitat offsets to critically assess the quality of constructed habitat for specialist anuran species that are associated with naturally acidic, oligotrophic waterbodies. This chapter also provides information that will inform

stakeholders if human intervention through anthropogenic waterbody construction can compensate for habitat loss of specialist wallum frogs.

3.3 Methods

Study Site Selection and Sampling Design

Waterbodies were located within and around Tyagarah Nature Reserve (NR) (28.6067°S, 153.5693°E) and the southern section of Bribie Island National Park (NP) (27.0732°S, 153.1774°E). Waterbodies were selected using Google Earth satellite imagery to cover the range of waterbodies available at each site. All waterbodies were surveyed twice, firstly between 18 October 2011 and 5 December 2011, and again between 7-21 February 2012. Thirteen natural (n = 5 Tyagarah, n = 8 Bribie Island) and twenty-five anthropogenic waterbodies (total n = 38) were surveyed for adult anurans, aquatic predators and tadpoles. Anthropogenic waterbodies consisted of roadside trenches/ditches (n = 4 Tyagarah; n = 5 Bribie Island), artificial ‘lakes’ (n = 1 Tyagarah; n = 8 Bribie Island) and old golf course waterbodies (n = 6 Tyagarah). With the exception of golf course waterbodies, all waterbodies contained natural vegetation around the vast majority (> 80%) of the waterbody perimeter.

Roadside trenches were waterbodies that had been constructed next to roads or firebreaks where earth material had been removed to aid construction of the road or firebreak. Artificial lakes were constructed waterbodies with the majority of the waterbody (> 80%) as open water with vegetation fringing the perimeter. Golf course waterbodies were constructed waterbodies within the golf course boundaries. Waterbodies within wallum heathland can be composed of heterogeneous vegetation types (refer to Chapter 2). Therefore, one large natural waterbody had multiple transects established a minimum of 300 m apart to ensure transect independence. Survey transects were established 5 m from the waterbody perimeter and ran parallel to the waterbody perimeter. Transect length varied between 50 – 100 m, depending on waterbody size.

Rainfall data was collected from the Australian Bureau of Meteorology Weather Station #058216 for Tyagarah NR and #040998 for Bribie Island NP. Sampling was conducted in 2011, with 246.9 millimetres (mm) recorded at Bribie Island and 373.4 mm recorded for Tyagarah NR during the three months prior to the date of the first survey. The 2012 surveys recorded 936.7 mm at Bribie Island and 405.2 mm for Tyagarah NR during the three months prior to the first survey date of the second survey. All waterbodies surveyed had water present during both the 2011 and 2012 surveys.

Diurnal aquatic predator traps (38cm length x 25cm width x 25cm height, with two square shaped entrances both 5cm x 5cm) were placed at the beginning, middle and end of each transect. Traps were baited with Orca Floating Fish Food Pellets and were left for approximately 30 minutes before being collected. Diurnal dip-netting for predatory fauna and tadpoles was conducted using a circular net, with an aperture of approximately 30 cm in diameter and mesh size less than 0.5 mm. Five ‘sweeps’ of the dip-net and measurements of pH and salinity were taken at 10 m intervals along each transect. Each ‘sweep’ included three water column levels (bottom, middle and top) to capture any variation in tadpole species richness or abundance that may occur along the water depth gradient (Heyer, 1973).

Water pH and salinity were measured using a TPS Aqua-CPA Conductivity-TDS-pH- Temperature Meter (version 1.2) approximately 15 cm below the surface of the water at every 10 m sampling point along the transect. Two 25 mL surface water samples were collected at the beginning and the end of each transect. Turbidity was measured using a HACH DREL 2000 Direct Reading Spectrophotometer within 24 hours of sample collection. Water chemistry measurements and samples were taken before dip-netting and trapping were conducted. Waterbody area was calculated using Google Earth satellite imagery.

Vegetation was categorized into five vegetation types; sedge, grass, lilly, Gahnia spp. and Melaleuca spp. Percent cover of each vegetation type was visually estimated using a 1 m x 1 m quadrat placed at each 10 m sampling interval.

Nocturnal anuran surveys were conducted by walking and visually searching the transect using a Princeton Apec headtorch. The species and the relative number of anurans seen for each species were recorded for each transect. For C. tinnula, numbers of individuals were determined using acoustic calls due to visual observations being absent or low at the survey transects. The vegetation type an individual anuran was first encountered on was also recorded. Substrate records were only conducted on the first survey.

Data Analysis

Relative abundances of L. olongburensis and L. fallax adults were higher during the 2011 survey. The cumulative number of L. olongburensis visually recorded during the first survey for each transect were divided by the transect length to obtain a relative abundance of L. olongburensis per meter. This was repeated for L. fallax, L. tyleri, Rhinella marina, and Limnodynastes peronii.

Incidental visual observations and frogs heard calling less than one meter outside the waterbody perimeter or calling within the waterbody during the first or second survey were recorded as occupying the waterbody. Tadpole species were considered present within a waterbody if they were recorded from any of the sampling points along transects.

Predatory fish caught in aquatic traps were used in data analysis due to low counts of predatory fish caught during dip-netting. The total number of predatory fish in each waterbody included individuals from the species Gambusia holbrooki (Eastern mosquito fish), Rhadinocentrus ornatus (Ornate rainbow fish), Hypseleotris galii (Firetail gudgeon), Hypseleotris compressa (Empire gudgeon) and Hypseleotris sp. (Midgleys carp gudgeon). The number of fish recorded for a given transect was divided by the number of aquatic traps used within the waterbody to obtain a relative fish abundance for each transect.

Salinity, turbidity and pH were averaged for each transect. Mean cover of each vegetation type for each transect was calculated by dividing the total percentage cover

of each vegetation type from each sampling point by the number of sampling points along the transect.

A Spearman Rank Correlation Test was performed in IBM SPSS Statistics Version 19 (SPSS, Inc., 2009, Chicago) to determine correlated variables. Highly correlated variables were considered to be variables that had a correlation coefficient value greater than or equal to 0.7 (sensu Babbitt et al., 2003; Garden et al., 2007). No variables were highly correlated and all variables were used in analyses.

A One-Way ANOVA with Tukey’s post-hoc analysis was used to determine significant differences in environmental variables between waterbody types. One way ANOVAs were also used to compare the relative abundance of L. olongburensis and L. fallax in the four different waterbody types.

The anuran community assemblage was analysed (using visual counts for L. olongburensis, L. fallax, L. peronii, R. marina and L. tyleri, and acoustic counts for C. tinnula) using non-metric Multidimensional Scaling (nMDS). These species were chosen as they were encountered in at least two waterbodies and were highest in relative abundance when compared with other species. Only waterbodies which contained one of these six species were used in nMDS analysis (n = 27). Bray-Curtis distance measures were used to determine waterbody similarities for the relative abundance of individual species. Four dimensions were used to minimise stress based on 999 random permutations to determine which environmental variables were significantly influencing the amphibian assemblage (Oksanen, 2011). Analyses were performed using the statistical program R (R-core Development Team 2011) using the vegan (version 2.04) package (Oksanen et al., 2012)

Twenty-eight models focusing on the influence of variables on the relative abundance of L. olongburensis and L. fallax were constructed a priori. To obtain the maximum number of predictor variables to use in each model, a generalised ‘rule of thumb’ of n/3 (where n = number of waterbodies sampled) (Crawley, 2007) was applied. Due to this rule failing to be adhered to by some models small sample Akaike’s Information

Criterion (AICc) was used to account for the number of model failing to meet this ‘rule’

(Burnham and Anderson, 2002). Predictor variables in L. olongburensis models were mean pH, average salinity, average turbidity, waterbody size, fish relative abundance, L. fallax relative abundance, % sedge cover, % Gahnia cover, % grass cover, % lily cover and % Melaleuca cover. Predictor variables in L. fallax models were the same as the L. olongburensis models, except L. fallax relative abundance was replaced with L. olongburensis relative abundance. Some variables (i.e. pH) were non-normally distributed. Therefore, Generalised Additive Mixed Effects models, using a Quasi- Poisson link function to account for overdispersion and location as a covariate, were used to determine the importance of the predictor variables on L. olongburensis and L. fallax relative abundance

Small sample Akaike’s Information Criterion was used for model selection, with the

‘best’ model having the lowest AICc value (Burnham and Anderson, 2002). To determine the ranking of the models, ΔAICci values were calculated, where higher

ΔAICi values indicated less accurate models for the given data (Burnham and Anderson,

2002; Johnson and Omland, 2004). If a model had a Δi ≤ 2, then there was considerable evidence that the model could be the “best” model, given the data (Johnson and

Omland, 2004). Models with a Δi ≤ 2-4 were considered to have moderate support for being the “best” model, given the data. Each model was assigned a model weight (wi) which was used to determine the “probability that model I is the best model for the observed data, given the candidate set of models” (Johnson and Omland, 2004). The closer the wi was to 1, the closer the model for the given data (Burnham and Anderson,

2002). To determine the relative importance of variables within models where Δi < 4 the wi values were summed from all models where the variable of interest occurred (Grueber et al. 2011). The closer the variable of interest was to 1 the higher the importance of the variable. All models were run in the freeware statistical package R (R Core Development Team, 2011) using the MuMIn (version 1.7.2) (Barton, 2012) and vegan (version 2.04) (Oksanen et al., 2012) packages.

3.4 Results

Waterbody Characteristics

Few waterbody characteristics varied among waterbody types (Table 3.1). Variables that were significantly different between waterbody types were pH (df = 36, F2/34= 5.36, p = 0.004), turbidity (df = 36, F2/34 = 3.925, p = 0.017) and waterbody size (df = 36,

F2/34 = 6.001, p = 0.002). Tukeys post hoc analysis revealed pH was significantly higher in artificial lakes than in natural waterbodies (p = 0.005) and roadside ditches (p = 0.011). Tukeys post hoc analysis also revealed that turbidity was significantly higher in natural waterbodies than in golf course waterbodies (p = 0.011) and natural waterbodies were significantly larger than road side ditches (p = 0.003) and golf course waterbodies (p = 0.024).

Anuran Assemblage / Occupancy

A total of 10 species were encountered; six recorded in natural waterbodies, six in artificial lakes, five in roadside ditches and six in golf course waterbodies (Figure 3.1). Litoria olongburensis and C. tinnula occupancy was highest within natural waterbodies while L. fallax, Litoria tyleri and R. marina occupancy was highest within artificial lakes (Figure 3.2). Litoria freycineti was only recorded from roadside ditches while L. gracilenta was only recorded from natural waterbodies. Additionally, one species of Uperolia sp. was recorded from one golf course waterbody. Litoria fallax was the only species to occupy all four waterbody types (Figures 3.1 and 3.2).

Water pH, salinity, % lily cover and % sedge cover significantly influenced the amphibian assemblage (Table 3.2). All species were separated on the nMDS plot, with L. olongburensis, C. tinnula and L. tyleri falling within close proximity of each other (Figure 3.3).

Table 3.1: Measured variable averages and ranges between the four waterbody types surveyed and for waterbodies with L. olongburensis and L. fallax. Variable Natural Roadside ditches Golf course Artificial lakes L. olongburensis L. fallax waterbodies waterbodies waterbodies waterbodies pH (average) 3.96 3.96 4.73 5.41 4.07 4.81 pH (range) 3.7 – 4.85 3.64 – 4.57 4.32 – 5.91 3.79 – 7.97 3.43 – 5.83 3.79 – 6.84

Salinity (ppm) 119.75 100.57 4684 153.89 90.8 109

Salinity (ppm) 50.22 – 108.67 73.56 – 206.67 40.6 – 27700 96.1 – 317.5 47.78 – 209.7 64.64 – 198.17 (range) Turbidity (FTU) 395 229 54 261.1 323 230

Turbidity (FTU) 150.5 – 889.33 21 – 217.33 30 – 100.33 16.7 – 724 21 – 889.33 16.6 – 724 (range) Area (m2) 22968 1196 337 11056 15574 8185 Area (m2) (range) 1726 – 51787 66 – 2666 750 – 5564 415 – 43351 376 – 51787 791 – 41129 % Sedge 55 19 5 22 52 28 % Sedge (range) 0 – 100 0 – 61 0 – 16 0 – 88 2 – 100 0 – 88 Fish abundance 2.13 2.09 3.06 2.9 3.07 3.42 Fish (range) 0 – 13 0 – 8.3 0.33 – 9 0 – 9.33 0 – 13 0 – 9.33

Figure 3.1: Amphibian species richness and species presence within each waterbody type. Colours/patterns indicate individual species.

Figure 3.2: Proportion of natural and anthropogenic waterbodies occupied for each recorded anuran species. Records are combined for both visual and acoustic records.

Table 3.2: Correlations (R2 values) between nMDS axis 1 and 2 and environmental variables influencing assemblage structure, with significant correlations (Pr (> r)) highlighted in bold. Variable NMDS1 NMDS2 R2 Pr (> r) pH 0.724 0.69 0.293 0.018

Salinity 0.046 0.999 0.258 0.033

Turbidity -0.988 -0.152 0.151 0.135

Area -0.998 0.057 0.091 0.316

Fish 0.706 0.708 0.002 0.975

Sedge -0.8 -0.599 0.251 0.039

Melaleuca -0.925 0.38 0.079 0.369

Fern -0.05 -0.999 0.105 0.241

Lily 0.941 0.339 0.311 0.016

Gahnia -0.022 -0.999 0.211 0.061

Grass -0.126 0.992 0.077 0.414

Litoria olongburensis and L. fallax relative abundance

Water chemistry variables differed among waterbodies containing L. olongburensis and L. fallax, with pH being 0.74 higher and sedge cover being 24% lower in waterbodies with L. fallax (Table 3.1). There was a significant difference between waterbody type for L. olongburnensis relative abundance (df = 36, F1/34 = 3.558, p = 0.025) but not for

L. fallax relative abundance (df = 36, F1/34 = 1.682, p = 0.19). However, L. fallax was only recorded from two natural waterbodies while the highest abundance of L. olongburensis was recorded from natural waterbodies (Figure 3.4). Additionally, no L. olongburensis were recorded from golf courses while the highest relative abundance of L. fallax was recorded from artificial lakes (Figure 3.4). 82

Litoria olongburensis were recorded perched on sedge (n = 153), Gahnia sp. (n = 14), Melaleuca sp. (n = 3), fern (n = 4) and grass (n = 2). Litoria fallax were recorded perched on sedges (n = 28), Gahnia sp. (n = 15), lilies (n = 14), Melaleuca sp. (n = 2) and ferns (n = 1).

Tadpoles of L. olongburensis were recorded in 11 natural waterbodies and two roadside ditches, which were adjacent to natural waterbodies. Tadpoles of L. fallax were only recorded from one natural waterbody adjacent to a roadside ditch. Tadpoles were only recorded during the summer survey in 2012 despite tadpole surveys being conducted during the 2011 and 2012 surveys.

One model for L. olongburensis relative abundance had a Δi ≤ 2 while no models had a

Δi ≤ 2-4. The weighting of the best model was 78.7%, indicating that the other models compared poorly (Table 3.3). This model contained % sedge cover and pH as key parameters. Litoria olongburensis abundance had a positive relationship with increasing % sedge cover and a negative relationship with increasing pH. Both % sedge cover and pH had high relative variable importance. However, pH had a confidence interval that intersected zero, diminishing the strength of the influence that this variable had on L. olongburensis relative abundance (Table 3.4).

Five models for L. fallax relative abundance had a Δi < 2 and five models had a Δi ≤ 2-

4. The combined weighting for models with Δi ≤ 2 was 61.8%. The combined weighting for models with Δi ≤ 2-4 was 28.6%. This gave a total weighting of 90.4% for all models with Δi ≤ 4, indicating that the other models compared poorly (Table 3.3).

Models with a Δi < 2 contained % sedge cover, L. olongburensis relative abundance, waterbody size, % fern density and % grass cover. All of these variables, with the exception of % grass cover, were negatively associated with L. fallax relative abundance. Models with a Δi ≤ 2-4 contained % Melaleuca cover, % Gahnia cover, % sedge cover, turbidity and predatory fish. Predatory fish and % Gahnia cover were the only variables within models where Δi ≤ 2-4 that was positively associated with L. fallax relative abundance. Sedge cover had the highest relative variable importance compared with the other variables within models where Δi ≤ 4. However, all variables

Model AICC Δi wi Litoria olongburensis % Sedge – pH 150.8 0.00 0.787 Litoria fallax (-) % Sedge 159.5 0.00 0.219 (-) L. olongburensis abundance 160.6 1.08 0.127 (-) Waterbody Size 160.9 1.46 0.105 (-) % Fern 161.4 1.88 0.085 % Grass 161.4 1.96 0.082 (-) % Melaleuca 161.8 2.32 0.069 (-) Turbidity 161.8 2.34 0.068 % Garnia 162.1 2.64 0.058 (-) % Sedge + % Garnia 162.3 2.85 0.053 Predatory Fish abundance 163.0 3.52 0.038

had a confidence interval that included zero, indicating reduced evidence that these variables did not have a strong influence on L. fallax relative abundance (Table 3.4).

3.5 Discussion

Anuran Assemblages

There was considerable overlap in the anuran species recorded from natural and anthropogenic waterbodies in wallum habitats of eastern coastal Australia. These results are similar to the majority of surveys conducted within the northern hemisphere (predominantly North America and Europe) that found anuran species richness to be higher or equal in artificial, restored waterbodies (reviewed in Brown et al., 2012). This is in contrast to a previous Australian study conducted in New South Wales comparing farm dams and natural waterbodies outside of wallum heathland areas, where natural 85

Table 3.4: Estimates for model averaged coefficients, standard error (SE), confidence interval (CI) and relative variable importance (RI) for each parameter in models where

Δi < 4 for L. olongburensis and L. fallax tadpole relative abundance. (+) indicates a positive relationship while (-) indicates a negative relationship between relative abundance and the model variable. Parameter Estimate SE CI RI Litoria olongburensis % Sedge 3.66 0.615 2.44, 4.86 0.94 pH -0.6 1.16 -2.88, 1.67 0.85 Litoria fallax (-) %Sedge -0.81 1 -2.78, 1.16 0.33 (-) L. olongburensis abundance -3.95 4.38 -12.53, 4.62 0.13 (-) Waterbody Size -0.00004 -0.00003 -0.0001, 0.00002 0.11 (-) % Fern -7.01 7.76 -22.22, 8.19 0.09 % Grass 0.39 3.22 -5.92, 6.7 0.11 (-) % Melaleuca -0.84 4.61 -9.87, 8.19 0.07 (-) Turbidity -0.0014 -0.0017 -0.004, 0.002 0.08 % Garnia 1.64 2.65 -3.55, 6.83 0.14 Predatory Fish abundance -0.023 -0.09 -0.15, 0.2 0.05

waterbodies had higher anuran species richness when compared with farm dams (Hazell et al., 2004). However, presence of an amphibian species at a waterbody is predominately associated with environmental variables, rather than the ‘status’ (i.e. natural / anthropogenic) of the waterbody (Mazerolle et al., 2005; Hagman and Shine 2006; Rannap, 2009).

The differences and similarities of these findings to previous studies are likely explained by the ability of individual species to respond to differences in the environmental variables that were found to be associated with the anuran assemblage (i.e. pH, % sedge cover, % lily cover). Vegetation variables significantly associated with the anuran assemblage are likely a function of particular species favouring certain varieties of vegetation. This was observed for L. olongburensis, which was found 86

predominantly on sedge species and supports past studies that found L. olongburensis prefers perching on certain sedge species (Shuker 2012). Intolerance to varying water chemistry levels by tadpoles of different anuran species explains the influence of water chemistry variables on the anuran assemblage, as adults have been shown to avoid depositing eggs in waters where water chemistry variables would likely be unfavourable for successful reproduction (Takahashi, 2007; Hamamura, 2008). For example, post Gosner stage 25 tadpoles of L. fallax failed to metamorphose when exposed to acidic waters (i.e. pH 3.5.), while acid-water adapted tadpoles species (i.e. C. tinnula) successfully metamorphosed in pH waters of 3.5, 4.5 and 5.0 (Meyer, 2004). This was further seen in Chapter two, where pH was associated with the relative abundance and occupancy of L. olongburensis tadpoles.

Intolerance to water chemistry variables only explains abundance and occupancy patterns for some species. Conversely, populations of species that are able to tolerate a wide range of water chemistry variable levels may be excluded from, or depressed within, waterbodies due to competition or predation. It has been hypothesised that competition may explain the parapatric distributions of L. fallax and L. olongburensis (Meyer et al., 2006), as shown by previous studies revealing competitive interactions between other amphibian species (Wiltshire and Bull, 1977; Twomey et al., 2008).

Surprisingly, the acid frog species L. freycineti was only recorded from roadside ditches. This species is infrequently encountered in natural waterbodies within coastal wallum systems; however, it is occasionally recorded from disturbed sites (i.e. drainage lines (Meyer et al., 2006), on roads away from wetlands (Hero pers. obs.) and fire trails near water (Simpkins pers. obs.)). These disturbed areas often have lower vegetation cover, which would likely increase detectability of this species when compared with natural waterbodies, where vegetation cover is often dense. Therefore, the absence of this species from natural waterbodies is potentially erroneous, complicating and the interpretation of natural habitat usage for this species. However, results presented in this chapter indicate that use of artificial lakes and golf course waterbodies by this species is low, possibly due to higher pH levels and competition from the non-acid frog species L. nasuta, which was recorded at these sites. Litoria nasuta has previously been proposed as a potential competitor of L. freycineti (Meyer et al., 2006).

Variables influencing L. olongburensis and L. fallax relative abundance

The variables most strongly associated with both L. olongburensis relative abundance was a high percent sedge cover and low pH. Sedge cover positively influenced L. olongburensis abundance with individuals predominantly found perching on sedge species. Amphibians may have coloration that assists with camouflage in their natural environment (Norris and Lowe, 1964; Toledo and Haddad, 2009). This may also occur with L. olongburensis, with sedge and dorsal colouration or patterning possibly aiding with camouflage (as suggested by Lowe and Hero, 2012). These results are consistent with a past survey that showed L. olongburensis to be habitat specialists using ecological niches within acidic (pH < 5) coastal wallum waterbodies containing sedges (Lewis and Goldingay, 2005; Shuker, unpublished data).

There were several variables that were associated with L. fallax relative abundance but it is difficult to identify discriminatory variables since those were within the top models had low relative variable importance and had confidence intervals that included zero. This suggests that a variable that is strongly influencing L. fallax relative abundance was not measured or that L. fallax is a generalist species, with the variables measured having the same relative influence on L. fallax relative abundance across the natural and modified landscape. Despite this, L. fallax numbers and occupancy within natural waterbodies (where pH was low and sedge cover high) were low and were highest within artificial lakes (where pH was high and sedge cover was intermediate). As mentioned previously, this is likely a result of L. fallax tadpole intolerance to low pH waters (Meyer, 2004). Additionally, these results indicate that L. fallax do not require waterbodies where sedge cover is high.

Implications for conservation

A range of waterbody types, both natural and anthropogenic/modified, are required to effectively conserve representative anuran assemblages within and around waterbodies that are naturally oligotrophic and acidic. However, conservation efforts should be focused primarily on natural waterbodies, where occupancy and reproduction of the

threatened, specialist species is highest. Conserving natural waterbodies would therefore increase the persistence of threatened species populations.

The presence of L. olongburensis in artificial lakes and roadside ditches and the presence of C. tinnula in artificial lakes and golf course waterbodies indicate that these waterbodies can provide habitat for adults of wallum associated threatened anuran species. These results corroborate recommendations that the establishment of artificial waterbodies with longer hydroperiods can be used to combat the effects of climate change in amphibian populations (Shoo et al., 2011). However, the low number of anthropogenic waterbodies with L. olongburensis tadpoles suggests that the majority of anthropogenic waterbodies may be unfavourable for breeding by this species. Subsequently, while providing potential refuge habitat, these waterbodies may be ecological traps that would not be able to sustain continuing L. olongburensis populations. Furthermore, the road ditches with L. olongburensis tadpoles were adjacent to natural waterbodies. Under these circumstances, road side ditches may be able to facilitate recruitment as well as connectivity between waterbodies.

With the exception of golf course waterbodies, the vast majority of waterbodies surveyed were surrounded by wallum heathland habitat. Therefore, results from this study may only be applicable to waterbodies where wallum heathland habitat is present. One ‘acid’ frog species (Litoria cooloolensis) has been located up to 1.3 kilometres away from natural waterbodies within undisturbed environments (Simpkins et al., 2011 - Appendix 1). This suggests dispersal of Litoria ‘acid’ frog species (i.e. L. olongburensis) could occur over equal distances and, provided that natural habitat is still intact, modified/disturbed waterbodies that are established away from natural waterbodies may still receive recruitment from natural waterbodies. Further study, where wallum heathland is not present around modified/disturbed waterbodies, is required to determine if ‘acid’ frogs can utilise waterbodies across disturbed landscapes.

Natural waterbodies where pH and sedge cover are within the ideal ranges for L. olongburensis should be given top priority for conservation. If construction or conservation of artificial waterbodies is undertaken for habitat loss compensation or in

an attempt to reduce the risks of climate change, then the variables that highly influence L. olongburensis (high sedge cover, low pH) must be attained.

It must be noted that habitat fragmentation and pond isolation may influence the assemblage of species within a waterbody. Pond isolation could not be measured for this study due to ponds being hidden from satellite images due to thick vegetation cover. Fragmentation was also difficult to measure because of this reason. Therefore, further study measuring pond isolation and pond fragmentation are needed to determine if these variables are influencing the amphibian assemblages within these ponds.

Acknowledgments

I would like to thank Clare Morrison, Evan Pickett, Jean-Marc Hero, Guy Castley, Katrin Lowe and Dianna Virki, who provided valuable comments on drafts of this chapter. We would also like to thank Edward Meyer for advice on pH data analysis and sampling techniques. We would like to thank Alan Kerr from the Bribie Island Environmental Protection Association who provided accommodation during fieldwork and Bayshore Resort in Byron Bay, New South Wales, Australia for giving us access to their land. We also thank Chris Touey, Tempe Parnell, Billy Ross, Matt Davies and Nick Clarke for assisting in data collection. Work conducted for this chapter was authorized by Griffith Animal Ethics number ENV/28/11/AEC, Queensland Department of Environment and Resource Management ECOACCESS permit number WITK10631812, and New South Wales Department of Environment Scientific License number S21717.

3.6 References

Babbitt, K.J., Baber, M.J., Tarr, T. L., 2003. Patterns of larval amphibian distribution along a wetland hydroperiod gradient. Canadian Journal of Zoology 81: 1539– 1552. Barton, K., 2012. MuMIn: Multi-model inference. R package, version 1.7.2. Page http://CRAN.R-project.org/package=MuMIn

Beebee, T.J.C., Griffiths, R.A., 2005. The amphibian decline crisis: A watershed for conservation biology? Biological Conservation 125: 271–285.

Brand, A. B., Snodgrass, J.W., 2010. Value of artificial habitats for amphibian reproduction in altered landscapes. Conservation Biology 24: 295–301.

Brown, D. J., Street, G.M., Nairn, R.W., Forstner, M.R.J., 2012. A place to call home: amphibian use of created and restored wetlands. Journal of Ecology 2012, doi:10.1155/2012/989872.

Burnham, K.P., Anderson, D.R., 2002. Model Selection and Multimodel Inference: A Practical Information-Theoretic Approach. Springer, New York.

Catterall, C.P., Kingston, M., 1993. Remnant Bushland of South East Queensland in the 1990’s: Its Distribution, Loss, Ecological Consequences, and Future Prospects. Institute of Applied Environmental Research, Griffith University and Brisbane City Council, Brisbane, Australia.

Chinathamby, K., Reina, R.D., Bailey, P.C.E., Lees, B.K., 2006. Effects of salinity on the survival, growth and development of tadpoles of the brown tree frog, Litoria ewingii. Australian Journal of Zoology 54: 97–105.

Christy, M.T., Dickman, C.R., 2002. Effects of salinity on tadpoles of the green and golden bell frog (Litoria aurea). Amphibia-Reptilia 23: 1–11.

Coaldrake, J.E., 1961. The eco-system of the coastal lowlands (Wallum) of southern Queensland. C.S.I.R.O Bulletin. No. 283.

Collins, J.P., Storfer, A., 2003. Global amphibian declines: sorting the hypotheses. Diversity and Distributions 9: 89–98.

Crawley, M.J., 2007. The R Book. Wiley, England.

Freda, J., Taylor, D.H., 1992. Behavioral response of amphibian larvae to acidic water. Journal of Herpetology 26: 429–433.

Garden, J.G., McAlpine, C.A., Possingham, H.P., 2010. Multi-scaled habitat considerations for conserving urban biodiversity: native reptiles and small mammals in Brisbane, Australia. Landscape Ecology 25: 1013–1028.

Garden, J.G., McAlpine, C.A., Possingham, H.P., Jones, D.N., 2007. Using multiple survey methods to detect terrestrial reptiles and mammals: what are the most successful and cost-efficient combinations? Wildlife Research 34: 218–227.

Griffith, S.J., Bale, C., Adam, P., 2008. Environmental correlates of coastal heathland and allied vegetation. Australian Journal of Botany 56: 512–526.

Griffith, S.J., Bale, C., Adam, P., Wilson, R., 2003. Wallum and related vegetation on the NSW north coast: description and phytosociological analysis. Cunninghamia 8, 202–252.

Grueber, C.E., Nakagawa, S., Laws, R.J., Jamieson, G., 2011. Multimodel inference in ecology and evolution: challenges and solutions. Journal of Evolutionary Biology 24: 699-711.

Haramura, T., 2008. Experimental test of spawning site selection by Buergeria japonica (Anura: Rhacophoridae) in response to salinity levels. Copeia 2008: 64-67.

Hagman, M., Shine, R. 2006. Spawning site selection by feral cane toads (Bufo marinus) at an invasion front in tropical Australia. Austral Ecology 31: 551-558.

Hazell, D. (2003). Frog ecology in modified Australian landscapes: a review. Wildlife Research 30: 193-205.

Hazell, D., Hero, J.-M., Lindenmayer, D., Cunningham, R., 2004. A comparison of constructed and natural habitat for frog conservation in an Australian agricultural landscape. Biological Conservation 119: 61–71.

Hecnar, S.J., M’Closkey, R.T., 1996. Amphibian species richness and distribution in relation to pond water chemistry in south-western Ontario, Canada. Freshwater Biology 36: 7-15.

Hecnar, S.J. , M’Closkey, R.T., 1997. The effects of predatory fish on amphibian species richness and distribution. Biological Conservation 79: 123–131.

Hero, J.-M., Magnusson, W.E., Rocha, C.F.D., Catterall, C.P., 2001. Antipredator defenses influence the distribution of amphibian prey species in the central Amazon rain forest. Biotropica 33: 131–141.

Heyer W.R., 1973 Ecological interactions of frog larvae at a seasonal tropical location in Thailand. Journal of Herpetology 7: 337-361.

Hines, H., Mahony, M. McDonald, K., 1999. An assessment of frog declines in wet subtropical Australia. Pp. 44- 63 In Campbell, A. (ed.) Declines and Disappearances of Australian Frogs. Environment Australia, Canberra.

Hines, H.B., Meyer, E.A., 2011. The frog fauna of Bribie Island: an annotated list and comparison with other Queensland dune islands. Proceedings of the Royal Society of Queensland 117: 261-274.

Ingram, G.J., Corben, C.J., 1975. The frog fauna of North Stradbroke Island, with comments on the “acid” frogs of the wallum. The Proceedings of the Royal Society of Queensland 86: 49–54.

Johnson, J. B., Omland, K.S., 2004. Model selection in ecology and evolution. Trends in Ecology & Evolution 19: 101–108.

Kats, L.B., Petranka, J.W., Sih, A., 1988. Antipredator defenses and the persistence of amphibian larvae with fishes. Ecology 69: 1865–1870.

Lemckert, L., Haywood, A., Brassil, T., Mahony, M. (2006). Correlations between frogs and pond attributes in central New South Wales, Australia: What makes a good pond? Applied Herpetology 3: 67-81.

Levin, N., Elron, E., Gasith, A., 2009. Decline of wetland ecosystems in the coastal plain of Israel during the 20th century: Implications for wetland conservation and management. Landscape and Urban Planning 92: 220–232.

Lewis, B. D., Goldingay, R.L., 2005. Population monitoring of the vulnerable wallum sedge frog (Litoria olongburensis) in north-eastern New South Wales. Australian Journal of Zoology 53: 185–194.

Lowe, K., Hero, J.-M., 2012. Sexual dimorphism and color polymorphism in the wallum sedge frog (Litoria olongburensis). Herpetological Review 43: 236–240.

Mazerolle, M.J., Desrochers, A., Rochefort, L. 2005. Landscape characteristics influence pond occupancy by frogs after accounting for detectability. Ecological Applications 15: 825-834.

Mazerolle, M. J., Poulin, M., Lavoie, C., Rochefort, L., Desrochers, A., Drolet, B., 2006. Animal and vegetation patterns in natural and man-made bog pools: implications for restoration. Freshwater Biology 51: 333–350.

Meyer, E., Hero, J-M., Shoo, L. and Lewis, B., 2006 National recovery plan for the wallum sedgefrog and other wallum-dependent frog species. Report to Department of the Environment and Water Resources, Canberra. Queensland Parks and Wildlife Service, Brisbane.

Mokany, A., Shine, R., 2002. Competition between tadpoles and mosquito larvae. Oecologia 135: 615–620.

Monello, R.J., Wright, R.G., 1999. Amphibian habitat preferences among artificial ponds in the Palouse region of northern Idaho. Journal of Herpetology 33: 298– 300.

Morin, P.J., 1986. Interactions between intraspecific competition and predation in an amphibian predator-prey system. Ecology 67: 713–720.

Oksanen, J., 2011. Multivariate analysis of ecological communities in R: vegan tutorial. http://ccoulufi/~jarioksa/opetus/metodi/vegantutorpdf, 20 September 2012.

Oksanen, J., Blanchet, F.G., Kindt, R., Legendre, P., Minchin, P.R., O’Hara, R.B., Simpson, G.L., Solymos, P., Stevens, M.H.H., Wagner, H., 2012. Vegan: Community Ecology Package, version 2.0-4. Page http://vegan.r-forge.r- project.org/, 20 September 2012.

Orizaola, G., BraÑa, F., 2003. Do predator chemical cues affect oviposition site selection in newts? Herpetological Journal 13: 189-193

Pierce, B.A., 1985. Acid tolerance in amphibians. BioScience 35: 239–243.

Rannap, R., Lõhmus, A., Briggs, L., 2009. Restoring ponds for amphibians: a success story. Hydrobiologia 634: 87–95.

Rios-López, N., 2008. Effects of increased salinity on tadpoles of two anurans from a Caribbean coastal wetland in relation to their natural abundance. Amphibia- Reptilia 29: 7–18.

Ruhí, A., Sebastian, O.S., Feo, C., Franch, M., Gascón, S., Richter-Boix, A., Boix, D., Llorente, G., 2012. Man-made Mediterranean temporary ponds as a tool for amphibian conservation. International Journal of Limnology 48: 81–93.

Shuker, J.D., Hero, J.-M., 2012. Perch substrate use by the threatened wallum sedge frog (Litoria olongburensis) in wetland habitats of mainland eastern Australia. Australian Journal of Zoology. doi.org/10.1071/ZO12030.

Shoo, L. P. et al., 2011. Engineering a future for amphibians under climate change. Journal of Applied Ecology 48: 487–492.

Simpkins, C. A., Meyer, E., Hero, J.-M., 2011. Long-range movement in the rare Cooloola sedgefrog Litoria cooloolensis. Australian Zoologist 35: 977–978.

Smallbone, LT., Luck, G.W. and Wassens, S. (2011). Anuran species in urban and scapes: relationships with biophysical, built environment ans socio-economic factors. Landscape and Urban Planning 101: 43-51.

Smith, M.J., Schreiber, E.S.G., Scroggie, M.P., Kohout, M., Ough, K., Potts, J., Lennie, R., Turnbill, D., Jin, C., Clancy, T., 2007. Associations between anuran tadpoles and salinity in a landscape mosaic of wetlands impacted by secondary salinisation. Freshwater Biology 52:75–84.

Sparling, D.W., 2010. Water-quality criteria for amphibians. Pp 105-120. in Dodd, Jr., C.K. (ed.) Amphibian Ecology and Conservation: A Handbook of Techniques. Oxford University Press, New York.

Strahan, R., 1957. The effect of salinity on the survival of larvae of Bufo melanostictus Schneider. Copeia 1957: 146–147.

Stuart, S.N., Chanson, J.S., Cox, N.A., Young, B.E., Rodrigues, A.S.L., Fischman, D.L., Waller, R.W., 2004. Status and trends of amphibian declines and extinctions worldwide. Science 306: 1783–1786.

Takahashi, M., 2007. Oviposition site selection: pesticide avoidance by gray treefrogs. Environmental Toxicology and Chemistry 26: 1476-1480.

Toledo, L.F., Haddad, C.F.B., 2009. Colors and some morphological traits as defensive mechanisms in anurans. International Journal of Zoology 2009: 1–12.

Turner, R.E., 1990. Landscape development and coastal wetland losses in the northern Gulf of Mexico. American Zoologist 30: 89–105.

Twomey, E., Morales, V., Summers, K., 2008. Evaluating condition-specific and asymmetric competition in a species-distribution context. Oikos 117: 1175–1184.

Vonesh, J.R., Kraus, J.M., Rosenberg, J.S., Chase, J.M., 2009. Predator effects on aquatic community assembly: disentangling the roles of habitat selection and post- colonization processes. Oikos 118: 1219–1229.

Wilbur, H.M., 1987. Regulation of structure in complex systems: experimental temporary pond communities. Ecology 68: 1437–1452.

Wiltshire, D.J., Bull, C.M., 1977. Potential competitive interactions between larvae of Pseudophryne bibroni and P. semimarmorata (Anura: Leptodactylidae). Australian Journal of Zoology 25: 449–454

Chapter 4 - Compensatory ponds provide poor habitats for the conservation of frogs associated with naturally oligotrophic, acidic environments

4.1 Abstract

Habitat loss is a key threatening process for many amphibians. Developers attempt to compensate for habitat loss by either restoring degraded habitat or artificially creating new habitats. Waterbodies constructed to aid in conservation of amphibians are quickly colonised by common species, however their utility for the conservation of threatened species has rarely been evaluated. Four compensatory and four established ponds in coastal wallum heathland of eastern Australia were surveyed over 20 months to determine if compensatory ponds were colonised by species from amphibian assemblages in adjacent established ponds. Additionally, the environmental variables associated with calling activity and relative abundance of two threatened species (Litoria olongburensis and Crinia tinnula) within the study area were determined. Established ponds contained different adult amphibian assemblages compared to compensatory ponds, with compensatory ponds being less suitable for the threatened amphibian species L. olongburensis and C. tinnula. Water pH significantly influenced amphibian assemblages within the landscape. Air temperature was the strongest variable positively influencing L. olongburensis calling activity, while daily and monthly rainfall were the strongest variables positively influencing calling activity by C. tinnula. Furthermore, pH and salinity were positively associated with the relative abundance of L. olongburensis while salinity, pH, water depth and minimum hydroperiod length were positively associated with the relative abundance of C. tinnula. Information on the optimal conditions for detecting L. olongburensis and C. tinnula and their preferred habitat characteristics need to be considered for restoring and constructing compensatory habitat for threatened amphibian species associated with naturally oligotrophic, acidic environments.

4.2 Introduction

Habitat loss is a key threatening process of global significance for the majority of vertebrate taxa (see Wilcove et al. 1998; Brooks et al. 2002; Cushman 2006). Compensatory, or artificial, habitats are often advocated and implemented when attempting to mitigate habitat loss or when restoring degraded habitat (Morris et al. 2006; Seaman 2007; reviewed in Brown et al. 2012). Additionally, compensatory habitats can be implemented as attempts to aid populations during extreme environmental events (i.e. extended droughts through climate change) (reviewed in Shoo et al. 2011). Compensatory habitats can be placed into two categories; ‘habitat creation’, where the type of habitat being created has not previously occurred within the construction area, or ‘habitat re-creation’, where the type of habitat being created is to mimic the former habitat (Morris et al. 2006). However, to be effective, compensatory habitats should mimic the original environment to replace the loss of the original environment. Unfortunately, there is little evidence to show that all environmental factors and abundance and richness of species can be accurately/successfully recreated (Morris et al. 2006; reviewed in Moreno-Mateos et al. 2012). Importantly, the time taken for compensatory habitats to reach a state that resembles the original environment can differ among habitat types (Morris et al. 2006; reviewed in Moreno-Mateos et al. 2012). This is of particular concern if compensatory habitats are to protect species that are already declining in an area.

Despite habitat re-creation being difficult, anthropogenic waterbodies designed to aid in conservation of amphibians, as well as those that are not, are still utilised by numerous amphibian species across different environments (Babbitt & Tanner 2000; Hazell et al. 2004; Barry et al. 2008; Rannap et al. 2009; Brand & Snodgrass 2010; reviewed in Brown et al. 2012). The majority of these studies focus on the northern hemisphere, with comparatively little research being conducted on southern hemisphere amphibians (reviewed in Brown et al. 2012), particularly within Australia. Therefore, the value of compensatory habitats for Australian amphibians is poorly understood and could impact on their successful design and implementation.

Numerous environmental variables may influence successful colonisation of anthropogenic waterbodies by amphibian species, including waterbody size (Babbitt & Tanner 2000), waterbody isolation, or habitat suitability (Lehtinen & Galatowitsch 2001). Additionally, environmental variables that influence amphibian assemblages within natural waterbodies (e.g. predators (Azevedo-Ramos et al. 1999; Hero et al. 2001) or hydroperiod (Babbitt et al. 2003; Babbitt 2005; Moreira et al. 2010)) may also influence amphibian usage of anthropogenic waterbodies. Environmental variables can be tolerated at different levels depending on the amphibian species (Gosner & Black 1957; reviewed in Pierce 1985; Christy & Dickman 2002). Therefore, the influence of environmental variables structuring amphibian assemblages within anthropogenic waterbodies will be species specific (Brown et al. 2012). Compensatory habitats may not provide suitable habitat for all species, particularly species that have a narrow tolerance to environmental factors.

With amphibians experiencing worldwide declines (Stuart et al. 2004) it is imperative that compensatory waterbodies are monitored to determine their efficiency in providing suitable habitats for amphibians. This chapter aimed to determine if newly constructed compensatory ponds contained similar frog assemblages when compared with established ponds over a two year study period within naturally oligotrophic, acidic environments. Compensatory habitats were hypothesised to be unsuitable for two threatened frog species (Litoria olongburensis and Crinia tinnula) that are associated with low pH waters (hereafter referred to as ‘acid’ frogs) and that habitat suitability would be a function of water chemistry. Furthermore, the study aimed to determine which environmental variables influenced the calling activity and relative abundance of L. olongburensis and C. tinnula to develop an optimised detection method for these two threatened species to assist in assessing habitat suitability of compensatory ponds.

4.3 Methods

Pond Description

Four compensatory ponds were constructed during the development of a four lane highway within south-eastern Queensland and north-eastern New South Wales,

Australia. Pond construction was aimed to compensate for removal of coastal wallum heathland that provided habitat for two threatened ‘acid’ frog species, Litoria olongburensis and Crinia tinnula. Four compensatory ponds (Ponds 1-4) were created within 30 meters of the roadside construction perimeter. Three established ponds (Ponds 6-8) had minimal anthropogenic disturbance occurring around their perimeter at the beginning of the survey period while one established pond (Pond 5) was isolated from native vegetation.

Survey Technique

Monthly surveying at the established and compensatory ponds commenced after completion of the compensatory ponds on the 20th September 2006 and finished on 10th May 2008. Surveys in the months of September 2007 and November 2007 were not conducted due to accessibility issues. Consequently, two additional surveys were conducted at the beginning and end of October 2007 to compensate for lost surveys. All ponds had a total of 20 surveys conducted over the study period, with the exception of compensatory pond 4 that had only 17 completed.

Acoustic and visual surveys were conducted for 10 minutes around the perimeter of each pond between 16hrs00min and 21hrs15min to maximise the probability of L. olongburensis being acoustically recorded (Hopkins unpublished data). During each survey, three to four acoustic playbacks of L. olongburensis and C. tinnula calls were conducted at randomly positioned points around each pond, with the random playback points alternating between surveys. Surveys were conducted, where possible, 24 hours (Hopkins unpublished data) to one week (Lewis & Goldingay 2005) after heavy rainfall to increase the probability of visually detecting L. olongburensis adults.

Abiotic Pond Characteristics

Abiotic pond characteristics (water pH, salinity and maximum depth) were measured for each pond before surveying commenced. Water chemistry parameters and water depth were measured from the deepest section of the ponds on the first survey. This point was used to measure water depth and water chemistry for all subsequent surveys. Water

chemistry was measured in the field until July 2007, after which water was obtained in the field and analysed back in the laboratory no more than 24 hours after the sample had been collected. Pond area estimates were calculated using Google Earth satellite photography by estimating the maximum length multiplied by maximum width of each waterbody. Distance to nearest vegetation patch was also calculated using Google Earth satellite photography. Air temperature was measured at each pond either at the beginning or during the survey period. Monthly, weekly and daily rainfall were obtained from the Australian Bureau of Meteorology Coolangatta Airport Rainfall Monitoring Station (Station number 040717), which is located adjacent to the study site.

Statistical Analysis

Water chemistry variables (pH and salinity) were combined separately for each pond and divided by the number of times water chemistry variables were surveyed for each individual pond. Minimum and maximum hydroperiod were determined by the lowest and highest number of consecutive surveys water was present within a pond. Air temperature was averaged from each pond for each survey period. The total number of adult frogs detected acoustically over the entire study period was combined for each species for each pond for analysis of variables influencing relative abundance of C. tinnula and L. olongburensis. For each survey period the number of calling C. tinnula and L. olongburensis encountered for each pond was used to determine the influence of environmental variables on calling activity of C. tinnula and L. olongburensis.

A one-way Analysis of Variance (ANOVA) was used to determine if there were significant differences in number of C. tinnula individuals between compensatory and established ponds. One-way ANOVAs were also used to determine if there was a significant difference in pond area, pH and salinity levels between compensatory and established ponds.

To determine which environmental variables were correlated, a Spearman Rank Correlation Test was performed in IBM SPSS Statistics Version 19 (SPSS, Inc., 2009, Chicago). Variables that had a correlation coefficient value greater than or equal to 0.7 (sensu Babbitt et al. 2003; Garden et al. 2007) were identified as being highly

100

correlated. No variables were highly correlated and thus all variables were used for data analysis.

Amphibian assemblages for each pond were represented using non-metric Multidimensional Scaling (NMDS). The Bray-Curtis distance measure was used to determine pond similarities for individual species relative abundance. Four dimensions were used to determine stress minimisation. Linear regression was used to determine significant environmental variables influencing the four MDS axes. Analysis was performed in the statistical program R (R-core Development Team 2011) using the vegan (version 2.03) package (Oksanen 2011).

Models focusing on the influence of environmental variables on calling activity and models focusing on the influence of environmental variables on relative abundance of L. olongburensis and C. tinnula were constructed a priori. Twenty-three models were constructed to model L. olongburensis and C. tinnula calling activity and relative abundance. Predictor variables included in L. olongburensis and C. tinnula calling activity models included water depth, air temperature and daily, weekly and monthly rainfall. Predictor variables included in L. olongburensis and C. tinnula relative abundance models included water depth, pH, salinity, pond size, distance to nearest vegetated habitat and maximum and minimum hydroperiod encountered during surveying. Generalised Additive Mixed-effects Models (GAMMs) were used to assess the importance of environmental variables on calling activity of L. olongburensis and C. tinnula (using a Poisson link function) with pond as a random effect and survey date as an auto-correlated structure. Generalised Linear Models (GLM) were used to assess the importance of environmental variables on relative abundance of L. olongburensis and C. tinnula (using a Poisson link function).

Akaike’s Information Criterion (AICc) was used to determine model ranking and selection. The best model was the model with the lowest AICc value (Burnham &

Anderson 2002). To determine the ranking of the models, Δi values were calculated, where higher Δi values indicated less accurate models for the given data (Burnham &

Anderson 2002; Johnson & Omland 2004). If a model had a Δi ≤ 2, then there was considerable evidence that the model could be the “best” model, given the data (Johnson

101

& Omland 2004). If a model had a Δi 2-4 then there was considered to be moderate evidence that the model could be the “best” model, given the data. Akaike Weights (wi) were assigned to each model to enable greater interpretation of the relative likelihood of a model (Burnham & Anderson 2002; Johnson & Omland 2004). The closer the wi was to 1, the stronger the model for the given data (Burnham & Anderson 2002). To determine the relative importance of variables within models where Δi < 4 the wi values were summed from all models where the variable of interest occurred (Grueber et al. 2011). The closer the variable of interest was to 1 the higher the importance of the variable. Models were run in R version 2.14.0 (R Core Development Team, 2011) using the ‘MuMIn’ (version 1.7.2) package.

4.4 Results

Waterbody Characteristics

Water pH was significantly higher (F1/95 = 39.513, p < 0.001) between compensatory ponds (4.00 – 7.54; average = 5.94 +/- 0.80) and established ponds (3.04 – 7.31; average

= 4.68 +/- 1.09). There was no significant difference in salinity levels (F1/77 = 1.283, p = 0.261) between compensatory ponds (0.02ppm – 0.62ppm, average = 0.16ppm +/- 0.17ppm) and established ponds (0.01ppm – 0.71ppm, average = 0.15ppm +/- 0.17ppm). Similarly, pond size did not differ significantly between compensatory and established ponds (F1/6 = 6.093, p = 0.195) with the largest and smallest compensatory pond being 206.09m2 and 82.51m2, respectively, while the largest and smallest established ponds were 1408.1m2 and 208.8m2, respectively.

Water was artificially added to the compensatory ponds in an attempt to ensure water was present within compensatory ponds, altering the ‘natural’ hydroperiods of the ponds. Artificial watering is believed to have resulted in water being present over a majority of the surveys for compensatory ponds one and three (17/20 surveys and 20/20 surveys). However, compensatory ponds two and four failed to maintain water for similar extended periods (11/20 surveys and 3/17 surveys). Established ponds did not receive artificial watering. Water length varied between established ponds, with water present during 17/20, 17/20, 11/20 and 12/20 surveys.

102

Minimum average air temperature recorded for the surveys was 14°C while maximum average temperature was 28°C. Minimum monthly rainfall was 9mm while maximum monthly rainfall was 279.2mm. Minimum weekly rainfall was 0.2mm while maximum weekly rainfall was 141mm. Minimum daily rainfall was 0mm while maximum daily rainfall was 24mm.

Compensatory and Established Pond Comparisons

Ten amphibian species (Crinia signifera, Crinia tinnula, Litoria fallax, Litoria nasuta, Litoria olongburensis, Litoria tyleri, Litoria peroni, Limnodynastes peronii, Opisthodon ornatus and Rhinella marina) were recorded from both established and compensatory ponds. An additional two species (Litoria rubella and Limnodynastes terraereginae) were only recorded at established ponds (Table 4.1). Individual species abundance varied between established and compensatory ponds, with threatened specialist species (L. olongburensis and C. tinnula) abundance being higher in established ponds (Table 4.1). Crinia tinnula abundance was significantly lower in compensatory waterbodies when compared with established waterbodies (F1/150 = 0.410, p = 0.003).

Non-metric multidimensional scaling using a Bray-Curtis distance measure was used to determine pond similarities for species assemblages between ponds (Figure 4.1). pH (49.7%) and depth (40.8%) explained the largest portion of MDS axis 1 (Table 4.2). However, only pH was significant in explaining the variation in species assemblages between ponds while depth was nearing significance (Table 4.2). The stress value associated with the NMDS plot was 0.001 (Figure 4.1).The influence of pH on the amphibian assemblage structure is clearly seen when plotted against the raw amphibian abundance data for each species (Figure 4.2).

103

Crinia signifera 2 0 1 0 3 0 2 1 2 5

Crinia tinnula* 18 5 16 5 44 32 51 64 28 175

Litoria fallax 26 5 8 1 40 2 6 0 3 11

Litoria nasuta 0 0 0 1 1 0 0 4 5 9

Litoria olongburensis* 1 0 0 0 1 9 7 15 21 52

Litoria tyleri 0 0 1 0 1 11 25 3 0 39

Litoria rubella 0 0 0 0 0 0 1 0 0 1

Limnodynastes peronii 10 12 0 0 22 0 0 6 9 15

Limnodynastes 0 0 0 0 0 4 3 1 1 9 terraereginae Opisthodon ornatus 1 0 0 0 1 0 1 0 0 1

Rhinella marina^ 3 12 8 1 23 0 6 1 1 8

104

105

Figure 4.2: Gradient analysis using average pH as a gradient with abudance of each species recorded across the survey period. N represents a natural pond while C represents a compensatory pond.

106

Table 4.2: Correlations to the MDS Axis 1-4 with variables playing a significant influence on assemblage structure highlighted in bold. A significant influence was considered a variable that had a p value less than 0.05. A* indicates significant variables while a # indicates a variable nearing significance (p = 0.052). Variable MDS Axis 1 MDS Axis 2 MDS Axis 3 MDS Axis 4 pH 0.4973* -0.0819 -0.1421 -0.1038 Salinity -0.1 -0.1365 -0.1353 0.0928 Pond Size 0.265 -0.0626 -0.1204 -0.02292 Hydroperiod -0.1574 -0.0566 0.1815 -0.0229 (min) Depth 0.4087# 0.1643 -0.0761 -0.0898 Distance to 0.1387 0.1991 -0.1658 -0.0652 nearest habitat Hydroperiod -0.0641 -0.0237 0.1247 -0.03957 (max)

Factors influencing relative abundance of L. olongburensis and C. tinnula

One model for L. olongburensis abundance had a Δi < 2. The weighting of this best model was 78.1%, suggesting that all other models compared poorly (Table 4.3). The best model contained pH and salinity (Table 4.3), with both these variables having strong relative importance (Table 4.4). One model also had a Δi 2-4 and a weighting of

16.4%. This model contained pH, salinity and depth. Combined, models with a Δi < 4 had a weighting of 94.5% (Table 4.3). However, relative variable importance confidence intervals for depth included zero, reducing evidence that this variable had a strong association with L. olongburensis abundance (Table 4.4).

Two models for C. tinnula abundance had a Δi < 2. The weighting of these best models was 97.6%, suggesting that the other models compared poorly (Table 4.3). The best models contained pH, salinity, depth and minimum hydroperiod length (Table 4.3). With the exception of pH, all of these variables had strong relative variable importance (Table 4.4). 107

L. olongburensis and C. tinnula calling activity

One model for L. olongburensis calling activity had a Δi < 2. The weighting of this best model was 86.1%, suggesting that the other models compared poorly (Table 4.3). The best model contained air temperature and monthly survey, which were both positively associated with calling activity of L. olongburensis (Table 4.3). Both of these variables had strong relative importance (Table 4.4), however, relative variable importance confidence intervals for these variables included zero, reducing the evidence that these variables were strongly associated with L. olongburensis calling. All other models had a

Δi > 4.

One model for C. tinnula calling activity had a Δi < 2. The weighting of this best model was 43.5% where monthly rainfall, daily rainfall and monthly survey were all positively associated with C. tinnula calling activity. Four models for C. tinnula calling activity had a Δi 2-4 with a combined weighting of 39.8%. The total weighting of all models with Δi < 4 was 83.3%, suggesting other models compared poorly. Variables within models with a Δi 2-4 were daily rainfall, weekly rainfall, monthly rainfall, water depth and survey date, with all variables except weekly rainfall being positively related to C. tinnula calling activity (Table 4.3). The relative importance of variables that were in models with a Δi 2-4 and not the model with a Δi < 2 were low when compared with variables in the model with a Δi < 2 (Table 4.4). Additionally, with the exception of monthly rainfall, all variables had relative variable importance confidence intervals that included zero, indicating reduced evidence that these variables have a strong association on C. tinnula calling activity (Table 4.4).

4.5 Discussion

Influence of environmental variables on the amphibian assemblage

Established and compensatory ponds supported different amphibian assemblages, with pH significantly associated with amphibian assemblages across the landscape. Frog assemblages within compensatory ponds were primarily influenced by high pH. Compensatory pond species exclusion or reduction in abundance from established

108

Model AICC ΔAICC w Litoria olongburensis calling activity Air Temperature + Monthly Survey 534.7 0.00 0.861 Litoria olongburensis abundance pH + salinity 38.4 0.00 0.781 pH + salinity + depth 41.5 3.13 0.164 Crinia tinnula calling activity Monthly Rainfall + Daily Rainfall + Monthly 363.1 0.00 0.435 Survey Monthly Rainfall + Daily Rainfall + Water 365.1 2.10 0.152 Depth + Monthly Survey Monthly Rainfall – Weekly Rainfall + Daily 366.1 2.99 0.097 Rainfall + Monthly Survey Water Depth + Monthly Survey 366.3 3.25 0.086 Daily Rainfall + Water Depth + Monthly 366.9 3.85 0.063 Survey Crinia tinnula abundance Depth + Salinity + Minimum Hydroperiod 67.6 0.00 0.698 Depth + Salinity + pH 69.5 1.85 0.278

109

Table 4.4: Model averaged coefficients for models where Δi < 4 for L. olongburensis and C. tinnula calling activity and relative abundance. Relative importance of each environmental predictor variable is displayed. Variable Estimate S.E. † Confidence Rel. var. Interval imp.‡ Litoria olongburensis calling activity Monthly Survey 0.102 0.055 -0.005, 0.209 1 Air Temperature 0.077 0.068 -0.056, 0.21 0.87 Litoria olongburensis abundance Salinity -14.204 0.436 -26.551, -1.859 1.00 pH -2.03 5.063 -3.155, -0.905 0.99 Depth 0.192 0.128 -0.163, 0.546 0.17 Crinia tinnula calling activity Monthly Survey 0.033 0.03 -0.027, 0.092 1 Daily Rainfall 0.015 0.009 -0.004, 0.034 0.78 Monthly Rainfall 0.004 0.001 0.001, 0.007 0.76 Water Depth 0.012 0.007 -0.002, 0.026 0.43 Weekly Rainfall -0.001 0.003 -0.006, 0.005 0.16 Crinia tinnula abundance Salinity -4.584 0.8449 -6.899, -2.269 1.00 Water Depth 0.085 0.0266 -1.112, -0.203 0.98 Minimum Hydroperiod -0.2172 0.0551 -0.37, -0.064 0.70 pH -0.6578 0.1755 -1.112, -0.203 0.30 † Standard Error ‡ Relative variable importance

110

ponds is likely due to species tolerances to differing levels of pH. For example, post- Gosner Stage 25 tadpoles of C. tinnula successfully metamorphose when exposed, under laboratory conditions, to pH levels of 3.5 (Meyer 2004). Additionally, adults of species comprising the majority of established pond assemblages (i.e. C. tinnula and L. olongburensis) successfully breed, in previous surveys, in waterbodies where minimum pH was between 3.2-3.4 (Meyer 2004). The majority of species occurring within compensatory ponds (i.e. L. fallax and L. peronii) were recorded from reproductive waterbodies where minimum pH was 5.0 (Meyer 2004). Therefore, the failure of compensatory pond species occurring in higher or equal numbers in the established ponds is likely due to intolerance to low pH levels. However, the low abundance of species comprising the established pond assemblage when compared with compensatory ponds is unlikely to be attributed to an intolerance of high pH waters as some species (C. tinnula and L. terraereginae) can successfully metamorphose when exposed to pH waters of 6.5 (Meyer 2004). The low abundance or exclusion of established pond species (i.e. L. olongburensis) from compensatory ponds could be contributed to competition from compensatory pond species. While no studies have been conducted to support this theory, it has been hypothesised that competition occurs between L. fallax and L. olongburensis (Meyer et al. 2006). Furthermore, studies on other amphibian species have revealed competitive interactions (Wiltshire & Bull 1977; Twomey et al. 2008), demonstrating competition can occur between amphibian species. Additionally, data from Chapter 2 indicates that competition may be restricting L. olongburensis at the upper limits of the ‘ideal’ pH range.

Variables influencing relative abundance of L. olongburensis and C. tinnula

Salinity and pH were both negatively associated with L. olongburensis and C. tinnula abundance while depth was positively associated with L. olongburensis and C. tinnula abundance. High salinity levels have been found to negatively influence survival of other amphibian tadpoles (Christy & Dickman 2002; Chinathamby et al. 2006; Rios- López 2008). The negative relationship with salinity and C. tinnula and L. olongburensis suggests that these species may be avoiding higher salinity levels encountered during this survey due to tadpole intolerance to high salinity levels.

111

However, salinity tolerance experiments on these species are required to fully discern this.

As discussed previously, the negative association between pH and L. olongburensis and C. tinnula abundance can be attributed to the these species having a broader tolerance to low pH conditions as opposed to sister species, which are less tolerate to these conditions. Furthermore, in high pH waters, competition from non-acid frog species is likely excluding these two acid-frog species from occurring in ponds with high pH while non-acid frog species tadpole intolerance is excluding amphibians within low pH ponds. This would allow acid tolerant species to occur in low pH ponds without the negative effect of competition.

Minimum hydroperiod was negatively associated with C. tinnula abundance. Crinia tinnula are often associated with ephemeral waterbodies (Meyer et al. 2006). Shorter hydroperiods would result in different predator assemblages when compared with waterbodies with longer hydroperiods (Welborn et al. 1996; Hero et al. 1998; Babbitt et al. 2003; Jocqué et al. 2007; Richter-Boix et al. 2007; Fernandes et al. 2010). Therefore, the negative association between C. tinnula and hydroperiod is likely due to the absence of more permanent aquatic tadpole predators, where C. tinnula would be unlikely to co-exist with. This can be seen in Chapter 2 of this thesis, that would inhibit coexistence of C. tinnula (see Chapter 2).

The positive association between water depth and L. olongburensis and C. tinnula is likely due to the species reproductive life history. Both species have tadpoles that occur in the water (Anstis 2007) and would need to deposit their eggs when risk of pond desiccation was low. For example, amphibian species may select their oviposition sites (Resetarits Jr & Wilbur 1989; Resetarits Jr 1996), with spawning occurring in ponds when water levels are at their peak (Goldberg et al. 2006), or in waterbodies that have the longest hydroperiod (Spieler & Linsenmair 1997).

The positive association between water depth and then negative association of minimum hydroperiod and C. tinnula could be perceived as ‘cancelling’ each other out. Dependant on the scenario, C. tinnula may be to seek a trade-off between pond

112

reliability and high risk of predation (increased water depth) and pond unreliability and safety (ephemeral/minimum hydroperiod).

Calling activity of L. olongburensis and C. tinnula

Air temperature was positively associated with L. olongburensis calling activity while previous daily and previous monthly rainfall were the variables with the highest relative importance for C. tinnula calling activity. Previous studies have shown that calling would not occur if air temperature fell below a certain threshold (Howard 1980; Saenz et al. 2006; Van Sluys et al. 2012), or where calling would be associated with, amongst other environmental variables, time of year and rainfall (Lemckert 2001; Saenz et al. 2006). The association between increased calling activity in L. olongburensis and air temperature can be related to amphibian biology. Amphibians are ectotherms and are likely to increase activity levels with increased temperature (Wells 2007). Daily and monthly rainfall increasing calling activity for C. tinnula is likely a reproductive response as the species attempts to breed when waters are peaking and risk of pond desiccation would be at its lowest.

Suitability of compensatory ponds for threatened species

The results presented here demonstrate these compensatory ponds are unable to mirror amphibian species assemblages of established ponds. Furthermore, these compensatory ponds were less effective in maintaining populations of the threatened, specialist species, C. tinnula and L. olongburensis, when compared with established ponds. Compensatory pond pH ranges were within the pH tolerance range (3.5 - 6.5 (Meyer 2004)) for C. tinnula tadpoles, suggesting that successful breeding could occur. However, the lack of any C. tinnula tadpole records within compensatory ponds (C. Simpkins, unpublished data) reduces the chances that successful breeding of C. tinnula occurred within compensatory ponds. Previous reports have noted C. tinnula association with disturbed waterbodies in wallum heathland (Meyer et al. 2006). However, this report did not mention if these populations successfully reproduced or if population numbers were the same in non-disturbed waterbodies (Meyer et al. 2006).

113

In this study, compensatory ponds were ineffective at maintaining populations of L. olongburensis and contained higher numbers of L. fallax when compared with established ponds. Sympatric co-existence between L. fallax and L. olongburensis rarely occurs, with L. fallax being proposed as a competitor of L. olongburensis (Meyer et al. 2006). Previous studies have found L. olongburensis adults occurring in ponds with pH between 3.8 – 4.6 (Hopkins unpublished data), 3.5 – 5.2 (Hero unpublished data) and 3.11 – 5.02 (Shuker unpublished data). Furthermore, the minimum pH of waterbodies within wallum where L. fallax have been found to occur is 5.0, with L. fallax eggs or tadpole death occurring when experimentally exposed to water of pH 3.5 (Meyer 2004). The mean pH of the compensatory waterbodies fell outside of the ranges where L. olongburensis have been detected while established pond pH averages fell within these ranges. Therefore, reduced numbers of L. olongburensis across compensatory ponds may be from L. fallax competition while the low abundance of L. fallax within established ponds is likely due to L. fallax larval intolerance to low pH waters.

Conclusions and implications for future establishment of compensatory habitats

The compensatory ponds surveyed were inadequate in their provision of suitable replacement habitats for wallum based amphibian species. The pH levels of the compensatory ponds are unlikely to exclude competitive species and/or their successful reproduction. Any attempts for future construction of compensatory habitats for threatened amphibian species need to take this into consideration and construct ponds where pH is kept within the optimal range of the species concerned. Additionally, pond depth needs to be within the deeper sections of the range detected for this survey. If construction of compensatory habitats for ‘acid’ frog species in eastern Australia is undertaken, post construction monitoring is recommended to determine if compensatory habitats are being utilised by C. tinnula and L. olongburensis. To maximise acoustic detection of C. tinnula and L. olongburensis within both compensatory and established ponds surveys should be conducted towards the upper limits of daily and/or monthly rainfall and the upper ranges of temperatures encountered during surveys. These results may only apply to L. olongburensis and C. tinnula populations within the study area and additional surveys of L. olongburensis and C. tinnula are required across the entire

114

distributional range of these species to determine if the variables within this study influence calling activity within other populations.

Acknowledgments

I would like to thank Jean-Marc Hero for providing me with the opportunity to work on the project that produced this chapter. I would also like to thank Stephen Lamb and Jon Shuker for helping with data collection. I would like to thank Queensland Main Roads for providing funding for data collection. I would like to thank Katrin Lowe, Jean-Marc Hero and Guy Castley for feedback on drafts of this chapter.

115

4.6 References

Anstis, M. (2002). Tadpoles of south-eastern Australia: a guide with keys. New Holland Publishers, Sydney.

Azevedo-Ramos, C., W. E. Magnusson, and P. Bayliss. 1999. Predation as the key factor structuring tadpole assemblages in a savanna area in Central Amazonia. Copeia 1999:22-33.

Babbitt, K. J. 2005. The relative importance of wetland size and hydroperiod for amphibians in southern New Hampshire, USA. Wetlands Ecology and Management 13:269-279.

Babbitt, K. J., M. J. Baber, and T. L. Tarr. 2003. Patterns of larval amphibian distribution along a wetland hydroperiod gradient. Canadian Journal of Zoology 81:1539-1552.

Babbitt, K. J., and G. W. Tanner. 2000. Use of temporary wetlands by anurans in a hydrologically modified landscape. Wetlands 20:313-322.

Barry, D. S., T. K. Pauley, and J. C. Maerz. 2008. Amphibian use of man-made pools on clear-cuts in the Allegheny Mountains of West Virginia, USA. Applied Herpetology 5:121-128.

Brand, A. B., and J. W. Snodgrass. 2010. Value of artificial habitats for amphibian reproduction in altered landscapes. Conservation Biology 24:295-301.

Brooks, T. M. et al. 2002. Habitat loss and extinction in the hotspots of biodiversity. Conservation Biology 16:909-923.

Brown, D. J., G. M. Street, R. W. Nairn, and M. R. J. Forstner. 2012. A place to call home: amphibian use of created and restored wetlands. Journal of Ecology 2012: doi:10.1155/2012/989872.

Burnham, K. P., and D. R. Anderson. 2002. Model selection and multimodel inference: a practical information-theoretic approach. Springer, New York.

Chinathamby, K., R. D. Reina, P. C. E. Bailey, and B. K. Lees. 2006. Effects of salinity on the survival, growth and development of tadpoles of the brown tree frog, Litoria ewingii. Australian Journal of Zoology 54:97-105.

Christy, M. T., and C. R. Dickman. 2002. Effects of salinity on tadpoles of the green and golden bell frog (Litoria aurea). Amphibia-Reptilia 23:1-11.

Cushman, S. 2006. Effects of habitat loss and fragmentation on amphibians: a review and prospectus. Biological Conservation 128:231-240.

Fernandes, I. M., F. A. Machado, and J. Penha. 2010. Spatial pattern of a fish assemblage in a seasonal tropical wetland: effects of habitat, herbaceous plant 116

biomass, water depth, and distance from species sources. Neotropical Ichthyology 8:289-298.

Garden, J. G., C. A. McAlpine, H. P. Possingham, and D. N. Jones. 2007. Using multiple survey methods to detect terrestrial reptiles and mammals: what are the most successful and cost-efficient combinations? Wildlife Research 34:218-227.

Goldberg, F. J., S. Quinzio, and M. Vaira. 2006. Oviposition-site selection by the toad Melanophryniscus rubriventris in an unpredictable environment in Argentina. Canadian Journal of Zoology 84:699-705.

Gosner, K. L., and I. H. Black. 1957. The effects of acidity on the development and hatching of New Jersey Frogs. Ecology 38:256-262.

Hazell, D., J.-M. Hero, D. Lindenmayer, and R. Cunningham. 2004. A comparison of constructed and natural habitat for frog conservation in an Australian agricultural landscape. Biological Conservation 119:61-71.

Hero, J.-M., C. Gascon, and W. E. Magnusson. 1998. Direct and indirect effects of predation on tadpole community structure in the Amazon rainforest. Australian Journal of Ecology 23:474-482.

Hero, J.-M., W. E. Magnusson, C. F. D. Rocha, and C. P. Catterall. 2001. Antipredator defenses influence the distribution of amphibian prey species in the Central Amazon Rain Forest. Biotropica 33:131-141.

Howard, R. D. 1980. Mating behaviour and mating success in woodfrogs, Rana sylvatica. Animal Behaviour 28:705-716.

Jocqué, M., T. Graham, and L. Brendonck. 2007. Local structuring factors of invertebrate communities in ephemeral freshwater rock pools and the inﬂuence of more permanent water bodies in the region. Hydrobiologia 592:271-280.

Johnson, J. B., and K. S. Omland. 2004. Model selection in ecology and evolution. Trends in Ecology & Evolution 19:101-108.

Lehtinen, R. M., and S. M. Galatowitsch. 2001. Colonization of restored wetlands by amphibians in Minnesota. The American Midland Naturalist 145:388-145.

Lemckert, F. 2001. The influence of micrometeorological factors on the calling activity of the frog Crinia signifera (Anura: Myobatrachidae). Australian Zoologist 31:625 - 631.

Lewis, B. D., and R. L. Goldingay. 2005. Population monitoring of the vulnerable wallum sedge frog (Litoria olongburensis) in north-eastern New South Wales. Australian Journal of Zoology 53:185-194.

117

Meyer, E. 2004. Acid adaptation and mechanisms for softwater acid tolerance in larvae of anuran species native to the “Wallum” of east Australia. PhD Thesis, University of Queensland.

Moreira, L. F. B., I. F. Machado, T. V. Garcia, and L. Maltchik. 2010. Factors influencing anuran distribution in coastal dune wetlands in southern Brazil. Journal of Natural History 44:1493-1507.

Moreno-Mateos, D., M. E. Power, F. A. Comı, and R. Yockteng. 2012. Structural and functional loss in restored wetland ecosystems. PLoS Biology 10: doi:10.1371/journal.pbio.1001247..

Morris, R. K. a., I. Alonso, R. G. Jefferson, and K. J. Kirby. 2006. The creation of compensatory habitat—can it secure sustainable development? Journal for Nature Conservation 14:106-116.

Pierce, B. A. 1985. Acid tolerance in amphibians. BioScience 35:239-243.

Rannap, R., A. Lõhmus, and L. Briggs. 2009. Restoring ponds for amphibians: a success story. Hydrobiologia 634:87-95.

Resetarits Jr, W. J. 1996. Oviposition site choice and life history evolution. American Zoologist 36:205-215.

Resetarits Jr, W. J., and H. M. Wilbur. 1989. Choice of oviposition site by Hyla chrysoscelis: role of predators and competitors. Ecology 70:220-228.

Richter-Boix, A., G. A. Llorente, and A. Montori. 2007. A comparative study of predator-induced phenotype in tadpoles across a pond permanency gradient. Hydrobiologia 583:43-56.

Rios-López, N. 2008. Effects of increased salinity on tadpoles of two anurans from a Caribbean coastal wetland in relation to their natural abundance. Amphibia- Reptilia 29:7-18.

Saenz, D., L. A. Fitzgerald, K. A. Baum, and R. N. Conner. 2006. Abiotic correlates of anuran calling phenology: the importance of rain, temperature, and season. Herpetological Monographs 20:64-82.

Seaman, W. 2007. Artificial habitats and the restoration of degraded marine ecosystems and fisheries. Hydrobiologia 580:143-155.

Shoo, L. P. et al. 2011. Engineering a future for amphibians under climate change. Journal of Applied Ecology 48:487-492. 118

Van Sluys, M., R. V. Marra, L. Boquimpani-Freitas, and C. F. D. Rocha. 2012. Environmental factors affecting calling behavior of sympatric frog species at an Atlantic Rain Forest Area, Southeastern Brazil. Journal of Herpetology 46:41-46.

Spieler, M., and K. E. Linsenmair. 1997. Choice of optimal oviposition sites by Hoplobatrachus occipitalis (Anura: Ranidae) in an unpredictable and patchy environment. Oecologia 109:184-199.

Stuart, S. N., J. S. Chanson, N. A. Cox, B. E. Young, A. S. L. Rodrigues, D. L. Fischman, and R. W. Waller. 2004. Status and trends of amphibian declines and extinctions worldwide. Science 306:1783-1786.

Twomey, E., V. Morales, and K. Summers. 2008. Evaluating condition-specific and asymmetric competition in a species-distribution context. Oikos 117:1175-1184.

Welborn, G. A., D. K. Skelly, and E. E. Werner. 1996. Mechanisms creating community structure across a freshwater habitat gradient. Annual Review of Ecology and Systematics 27:337-363.

Wilcove, D. S., D. Rothstein, J. Dubow, A. Phillips, and E. Losos. 1998. Quantifying threats to imperiled species in the United States. BioScience 48:607-615.

Wiltshire, D. J., and C. M. Bull. 1977. Potential competitive interactions between larvae of Pseudophryne bibroni and P. semimarmorata (Anura: Leptodactylidae). Australian Journal of Zoology 25:449-454.

119

Chapter 5 - Comparison of predation rates between the introduced mosquito fish (Gambusia holbrooki) and native aquatic predators on L. olongburensis, L. fallax and Limnodynastes peronii tadpoles

5.1 Abstract

Predation by aquatic predators is often an underlying factor in structuring tadpole assemblages. Predation effects may increase when tadpole communities are exposed to ‘exotic’ predators, as evolved anti-predator strategies may be ineffective. Few studies have compared predation rates of introduced and native tadpole predators on tadpole species. This chapter tested and compared predator effectiveness of an introduced predator (Gambusia holbrooki) with one native fish species (Hypseleotris galii) and four native aquatic invertebrates (Belostomatidae (water scorpion), Ashnidae (dragonfly larvae), Zygoptera (mayfly larvae) and Cherax sp. (crayfish)) on ‘small’ and ‘large’ tadpoles of Limnodynastes peronii (Striped Marsh Frog), Litoria fallax (Eastern Sedge Frog) and Litoria olongburensis (Wallum Sedge Frog). Predator effectiveness was dependent on the tadpole and predator species; Belostomatidae were the most effective predators within single prey experiments while Cherax sp. were the most effective predators within multiple prey experiments. Surprisingly, in the single tadpole experiments, the native fish H. galii did not attack any tadpoles. However, the introduced predator G. holbrooki attacked 3 of 12 tadpoles. Within the multiple prey experiments, H. galii consumed an average of 4.87 tadpoles while G. holbrooki consumed an average of 2.23 tadpoles. However, the predator species that consumed the highest average number of tadpoles differed between tadpole species. These findings indicate that while G. holbrooki were either less effective or equal predators of tadpole. However, these results indicate that native predatory species may have greater potential to influence the tadpole assemblage than G. holbrooki.

120

5.2 Introduction

Predation by aquatic predators is often an underlying factor structuring tadpole assemblages (Hero et al., 1998; Hero et al., 2001; Vonesh et al., 2009), with aquatic tadpole predators including, but not limited to, predatory fish (Hero et al., 2001; Baber and Babbitt, 2003; Gregoire and Gunzburger, 2008; Nelson et al., 2011) crayfish (Axelsson et al., 1997; Stoneham, 2001), Odonate larvae (Jara, 2008; Álvarez and Nicieza, 2009) and predacious tadpoles (Heyer et al., 1975; Wells, 2007).

Tadpole predation risk is determined by a number of factors, but primarily predator assemblages and predator abundance, which varies among waterbodies (Woodward, 1983; Hero et al., 1998). Furthermore, a predators ability to exclude or reduce tadpole populations will also be influenced by the tadpoles anti-predator strategies, which can include, but are not limited to, unpalatability/chemical defences (Kats et al., 1988; Crossland, 2001; Hero et al., 2001; Gunzburger and Travis, 2005), behavioural avoidance (Skelly, 1994; Saidapur et al., 2009; Smith and Awan, 2009) or morphological adaptations (Hecnar and M'Closkey, 1997; McCollum and Leimberger, 1997; Touchon and Warkentin, 2008).

Anti-predator strategies may only be effective against specific predators. Unpalatability, for example, may work at deterring predatory fish species but be ineffective against some aquatic invertebrate predators (Azevedo-Ramos et al., 1992; Hero et al., 2001). Additionally, certain strategies will be more effective at particular stages of development, with palatability found to decrease with development in Rhinella marina (previously Bufo marinus) tadpoles (Lawler and Hero, 1997). Increased tadpole body size can decrease the risk of attacks from predatory species (Jara, 2008), thereby influencing tadpole predation rates.

The introduction of non-native predators into waterbodies has been hypothesised as contributing towards amphibian declines (Morgan and Buttemer, 1996; Gillespie and Hero, 1999; Kats and Ferrer, 2003; Pyke, 2008), due to tadpole species not evolving adequate anti-predator defences to prevent predation by the introduced predator (Kats and Ferrer, 121

2003). Hence, determining the effectiveness of an introduced predator on prey populations cannot be fully discerned without comparing the influence of native predators on the prey population. To my knowledge, only one study has conducted preliminary experiments to determine if native aquatic predators of tadpoles have the same predation levels as G. holbrooki (Pyke and White, 2000), an introduced fish species in several countries. This study found that predation levels on Litoria aurea tadpoles were lower for native predators when compared with G. holbrooki (Pyke and White, 2000).

To expand on this previous work this chapter examines the predation rates for native aquatic predators (Cherax sp., Hypseleotris galii, Belostomatidae, Aeshnidae and Zygoptera) and one introduced aquatic predator (G. holbrooki) on larvae of a threatened ‘acid’ tadpole species (Litoria olongburensis) and a ‘non-acid’ species (Litoria fallax), which has been identified as a potential competitor of L. olongburensis (Meyer et al., 2006). We also tested predator effectiveness on tadpoles of Limnodynastes peronii, a widespread species known to occur within tadpole assemblages containing both L. fallax and L. olongburensis. Results of this study can aid in determining the influence of G. holbrooki on amphibian communities in waterbodies where G. holbrooki have been introduced. It should be noted that results obtained for this study are from an artificial environment and interpretation of these results into the ‘real world’ environment needs to be extended with future studies.

5.3 Methods

Collection of egg and larvae

Fertilised eggs from L. fallax were obtained by placing one adult male and one adult female frog into a plastic zip-lock bag (22 cm x 25 cm) that contained approximately 500mL of water. Water was obtained from the waterbody where the adults were caught. Sedges or other forms of aquatic vegetation were also placed in the bag. The pair of frogs were then left overnight to facilitate spawning. This was repeated several times using different individual frogs until a sufficient number of eggs had been obtained to conduct predation 122

experiments. Several clutches of eggs from L. fallax were obtained from an ephemeral wetland (-27.965712°, 153.379338°) on the 23rd September 2011 and the 18th January 2012. Two additional L. fallax egg clutches were also obtained from a permanent wetland at Currumbin Wildlife Sanctuary, Currumbin, Queensland, Australia (-28.139091°, 153.484030°) on the 15th November 2011. Eggs from L. peronii were obtained from three foam egg masses found in a permanent waterbody at Musgrave Park, Southport, Queensland, Australia (-27.957785°, 153.394208°) on the 25th August 2011. Eggs were placed into an eight litre plastic bucket with water from the waterbody where their parents were captured or from where eggs were collected. Eggs hatched in the bucket and were fed frozen lettuce until tadpoles reached lengths between 8-13mm (hereafter referred to as ‘small’) and 16-27mm (hereafter referred to as ‘large’).

Several attempts using the described methods for obtaining eggs from L. fallax failed to produce any L. olongburensis adult pairs spawning eggs. Therefore, tadpoles were dipnetted from a semi-permanent waterbody at Bribie Island National Park (-27.074697°, 153.178290°). Tadpoles from L. olongburensis were between 21-34 mm for experiments containing multiple tadpoles and 8-38 mm for experiments containing one tadpole.

Predation Experiments

White plastic containers (43 cm length x 34 cm width x 17 cm height) holding approximately eight litres of pond water were used for predation experiments. A secure fibreglass mesh lid (mesh size 0.1cm x 0.1cm) was placed over each container to ensure no interference from ‘outside’ fauna occurred and to ensure that ‘climbing’ predators (i.e. crayfish) could not escape. Water was collected from the surface of the waterbody to reduce the amount of detritus that was collected. Predator-prey experiments were conducted next to waterbodies where the predators were collected to minimise the risks of transporting predators and to expose experiments to natural climatic conditions.

Potential predators were captured using aquatic traps (38cm length x 25cm width x 25cm height, with two square shaped entrances both 5cm x 5cm) and were baited with 123

approximately 20 grams of floating fish food pellets. Aquatic invertebrates were caught using dip-netting techniques. Individual predators and tadpoles were measured to the nearest millimetre using a 30 centimetre ruler before placement into experimental containers. Fish were measured from the mouth to the tail tip, crayfish from the rostrum to the telson, and tadpoles from the mouth to the tip of the flagellum.

Individual L. olongburensis tadpole predation experiments

Due to the difficulty in finding tadpoles of L. olongburensis, only one L. olongburensis tadpole and one predator were placed into a plastic container at 1200hrs and signs of predation were checked at 1800hrs the following day. This allowed for 30 hours of predator-prey interaction time. These individual interactions were replicated for each predator, with replication being dependant on the number of predators captured (Figure 1). A maximum of 20 experiments were conducted at a time. Variation in time between predator capture and placement within the containers varied between one to two hours. No tadpoles or predators were re-used in any experiments and were released at the site of capture at the end of each experiment. Tadpoles were categorised as attacked if visual signs of predatory damage were visible on the flagellum and categorised as consumed/eaten if no tadpole was present at the end of the experiment.

The percentage of consumed tadpoles was used to determine predator effectiveness in single tadpole experiments. Percentages for each predatory species was obtained by adding the number of experiments where the tadpole was consumed and dividing by the total number of experiments used for the predatory species in question.

Multiple tadpole experiments

Aquatic predators were placed into white plastic containers and starved for twelve hours prior to the addition of 10 tadpoles from a single anuran species into each container. The number of tadpoles surviving were checked and recorded every 12 hours over a 48 hour period. Tadpoles were considered to be consumed if a tadpole was absent. Consumed 124

tadpoles were replaced with new tadpoles every 12 hours. The total number of tadpoles consumed within one container over the 48 hour period was considered as one replicate when undertaking data analysis. This was replicated for each tadpole species and predator species, with the number of experiments dependent on the availability of tadpoles and predators (Table 5.1). No tadpoles or predators were re-used in experiments and all surviving tadpoles were released at the site of capture at the end of each experiment.

Predator size can influence predation rates on tadpoles (Webb and Joss, 1997). Therefore, for multiple-prey experiments, a One-Way ANOVA was used to determine if there were significant differences in size of predator species between tadpole species. Tukey’s honest signiﬁcant difference (HSD) post hoc t-tests were then used to determine which tadpole species were influencing the significant differences.

An ANCOVA was used to examine the relationship between the total number of tadpoles consumed and the independent variables; tadpole species, predator species and predator size as a covariate. ANCOVAs were also used to determine if the number of tadpoles that were predated upon were significantly different among predator species.

Table 5.1: Number of experiments conducted for each tadpole predator species for multiple prey experiments. Tadpole Type Cherax sp. H. galii G. holbrooki Total

Litoria olongburensis 4 8 6 18 Litoria fallax small 6 10 8 24 Litoria fallax large 4 3 5 12 Limnodynastes peronii 10 10 11 31 Total 24 31 30 85

125

5.4 Results

Individual Tadpole Experiments

The efficiency of predators varied amongst taxa (Figure 5.1). The predacious beetles, in the family Belostomatidae, were the most effective predators, consuming 90% (9/10 tadpoles) of L. olongburensis tadpoles offered to them. The next most effective predators were members of the Aeshnidae family and members within the Cherax genus, which consumed 27% (3/11 tadpoles) and 10% (1/10 tadpoles) of L. olongburensis tadpoles offered to them, respectively. Individuals of the species G. holbrooki attacked, but did not consume, 25% (3/12 tadpoles) of L. olongburensis tadpoles offered to them. All tadpoles attacked by G. holbrooki were alive at the end of the study. Individuals of the native fish species H. galii and larvae of the sub-order Zygoperta neither consumed nor attacked any L. olongburensis tadpoles (Figure 5.1).

Multiple Prey Experiments

There was a significant difference in size of Cherax sp. (df = 23, F3/20=18.025, p = 0.000) and H. galii (df = 30, F3/27 =3.381, p = 0.033) used among tadpole species. Tukeys post-hoc t-tests revealed that the significant differences for Cherax sp. size were between L. olongburensis treatments and all other tadpole species (L. fallax large, (M = 34, 95% CI [20.21, 47.79], p =0.000); L. fallax small, (M = 28.25, 95% CI [13.14, 43.36], p =0.000); Lim. peronii, (M = 28.5, 95% CI [15.86, 41.14], p =0.000). Tukeys post-hoc t-tests for H. galii size revealed that the size difference between L. olongburensis and Lim. peronii (M = 8.8, 95% CI [-18.14, 0.54], p =0.7) and between L. olongburensis and ‘small’ L. fallax (M = 9.2, 95% CI [-18.54, 0.14], p =0.55).

There was a significant interaction between tadpole species and predator species (df = 6, F = 2.8, p = 0.017) for the total number of consumed tadpoles. Overall, Cherax sp. consumed the largest number of tadpoles while G. holbrooki consumed the lowest number of tadpoles (Table 5.2). 126

There was a significant difference in the number of tadpoles consumed between tadpole species for Cherax sp. (df = 3, F3/20 = 4.232, p =0.019) and H. galii (df = 3, F3/27 = 7.38, p = 0.001), with Cherax sp. and H. galii consuming the highest number of Lim. peronii tadpoles and the lowest number of L. olongburensis tadpoles (Figure 5.2 and Table 5.2). Despite G. holbrooki consuming the highest number of tadpoles in large L. fallax and Lim. peronii experiments, there were no significant differences in the number of tadpoles consumed between tadpole species for G. holbrooki (df = 3, F3/26 = 1.36, p = 0.279). With the exception of small L. fallax tadpoles, G. holbrooki consumed, on average, fewer tadpoles when compared with the other two native predatory species. 127

Table 5.2: Average number of tadpoles consumed for each predator species during multiple prey experiments. Tadpole Type Cherax sp. H. galii G. holbrooki

Litoria olongburensis 4.5(+/- 5.8) 0.6 (+/- 0.89) 0.55 (+/- 1.66) Litoria fallax small 8 (+/- 2.28) 3.2 (+/- 2.4) 4.13 (+/- 1.73) Litoria fallax large 7.75 (+/- 4.65) 2.67 (+/- 1.53) 0.8 (+/- 0.45) Limnodynastes peronii 14 (+/- 4.32) 10.8 (+/- 8.68) 2.27 (+/- 5.29) Total Consumed 9.88 (+/- 5.46) 4.87 (+/- 6.57) 2.23 (+/ - 3.59)

There was a significant difference in the number of Lim. peronii tadpoles consumed (df = 2,

F2/28 = 4.37, p = 0.023) and the number of small L. fallax consumed (df = 2, F2/21 = 11.74, p = 0.000) among predator species. Cherax sp. consumed the highest number of tadpoles in both Lim. peronii and small L. fallax experiments (Figure 5.2 and Table 5.2). There was no significant difference in the number of large L. fallax tadpoles consumed (df = 2, F2/15=

3.552, p = 0.079) or the number of L. olongburensis (df = 2, F2/15=0.746, p = 0.492) tadpoles consumed among predator species.

5.5 Discussion

This study is the first to quantify differences in Australian tadpole predation levels between G. holbrooki and native aquatic predators using an experimental approach. As expected, predation rates varied between tadpole species and predatory species, with G. holbrooki predation levels being approximately equal to or less then predation levels of native predatory species.

128

129

Gambusia holbrooki predation rates

These results are the first, to my knowledge, to show that G. holbrooki are either less or equally effective as tadpole predators when compared with native predatory species. Only one other study has provided preliminary comparative data for G. holbrooki predation levels with four Australian native predatory fish species (Pyke and White, 2000). Pyke and White (2000) showed that native fish predators had lower tadpole predation rates when compared with G. holbrooki. Results from the current study contrast with these as native fish within the multiple prey experiments consumed a lower number of tadpoles when compared with G. holbrooki. Furthermore, crayfish and some aquatic invertebrate predators had higher consumption rates when compared with G. holbrooki in both multiple and single prey experiments.

Lower tadpole consumption of L. olongburensis and large L. fallax tadpoles by G. holbrooki and H. galii when compared with the other two, smaller, tadpole categories may be explained by physical limitations of the predator (i.e. mouth gape limitation) or unpalatability, which can increase in older tadpoles (Lawler and Hero, 1997). Mouth-gape limitations by these two predatory species would result in tadpole consumption being restricted to smaller tadpoles. If mouth-gape limitations were influencing predation rates on larger tadpoles then G. holbrooki may have used the observed tail-nipping to reduce tadpole movement in an attempt to kill the tadpole to allow for easier consumption, regardless of tadpole size (Baber 2001, cited in Baber and Babbitt 2003). Similar tail-nipping attacks have been observed where G. holbrooki attacked, but did not kill, L. aurea tadpoles (Webb 1994, cited in Morgan and Buttemer 1996; Pyke and White 2000). Furthermore, tadpole body size of Crinia signifera and L. peronii tadpoles in past experiments did not influence predation rates of G. holbrooki (Webb 1994, cited on Morgan and Buttemer 1996).

Lower predation rates of small L. peronii tadpoles when compared with small L. fallax tadpoles by G. holbrooki are unlikely to be due to mouth-gape limitations as the sizes of tadpoles within these two tadpole groups were similar. Furthermore, it is unlikely that unpalatability towards fish is the cause of low predation rates, as the other fish species (H. 130

galii) used in the experiment had relatively high predation levels when compared with G. holbrooki. While not tested, other anti-predator strategies like behavioural avoidance (Skelly, 1994; Saidapur et al., 2009; Smith and Awan, 2009) or other morphological adaptations (Hecnar and M'Closkey, 1997; McCollum and Leimberger, 1997; Touchon Warkentin, 2008) are likely to be the cause of lower predation rates by G. holbrooki on Lim. peronii tadpoles.

Predator Effectiveness

The primary consumers of tadpoles within single prey experiments were aquatic invertebrate insects within the Belostomatidae and Aeshnidae families. These results support past studies that found Belostomatidae and Aeshnidae were effective predators of tadpoles (reviewed in Wells, 2007), with tadpoles contributing a large portion of Belostomatidae diet (Ohba and Nakasuji, 2006). Additionally, members within the Aeshnidae family can exclude or reduce tadpole populations from waterbodies (Hero et al., 2001; Stav et al., 2007) and have been known to predate on tadpole species that are unpalatable to fish (Hero et al., 2001).

There was no predation of L. olongburensis tadpoles by individuals in the sub-order Zygoptera. A literature search on predators of amphibian larvae (conducted by Wells, 2007) showed no publications recording Zygoptera as a predator of tadpoles. These results further validate these findings and indicate that Zygoptera are ineffective predators of L. olongburensis tadpoles.

A higher number of tadpoles were consumed by Cherax sp. when compared with other predators in tadpole experiments. Crayfish have been described as both active (Wells, 2007) and sit and wait (Renai and Gherardi, 2004) predators that have the potential to influence tadpole assemblages (Dorn and Wojdak, 2004). The high level of tadpole predation indicates that Cherax sp. have a greater ability to influence the tadpole community when compared with other predatory species used in predator experiments (i.e. G. holbrooki, H. galii). Furthermore, crayfish were shown to leave killed 131

bufonid/unpalatable tadpoles (Axelsson et al., 1997), suggesting the differing predation levels of Cherax sp. between tadpole species may give an indication of tadpole palatability.

The lower number of L. olongburensis tadpoles consumed by Cherax sp. when compared with other tadpole categories in multiple prey experiments may be due to significantly larger Cherax sp. used in L. olongburensis tadpole experiments. Crayfish are polytrophic omnivores that will consume flora or fauna material (Axelsson et al., 1997; Verhoef et al., 1998; Furse and Wild, 2004). However, adult crayfish are often found with high levels of detritus and plants in their gut content while juvenile crayfish feed predominately on invertebrates (Nyström, 2002). This would explain patterns observed in the current study, with smaller crayfish being younger than larger crayfish and having higher predation levels on fauna, possibly due to of higher protein requirements compared with their adult counterparts.

The number of experiments where Cherax sp. and H. galii consumed tadpoles was lower in experiments with single L. olongburensis tadpoles than those with multiple L. olongburensis tadpoles. An increase in prey abundance and longer experimental time for multiple prey experiments would allow for a higher likelihood of predator-prey interactions and, possibly, allow for higher predation levels. Alternatively, the starvation period in predator consumption rate experiments may have contributed to increased predation levels.

Conservation Implications

This study indicates that G. holbrooki have the potential to be just as effective as tadpole when compared with native predatory species. When present, G. holbrooki are often the most abundant fish species (reviewed in Pyke, 2008) and this would therefore increase the frequency of predator-prey interactions and potentially influence the abundance of amphibians at the waterbody. This is supported by past studies, that have shown a negative relationship with amphibian abundance and G. holbrooki presence (Webb and Joss, 1997). 132

Amphibian species richness, however, is unlikely to be influenced by presence of G. holobrooki, with past field studies finding that presence of L. aurea tadpoles (Hamer et al., 2002) and adults of other Australian native amphibian species (Reynolds, 2009) were not influenced by the presence of G. holbrooki. Hence, waterbodies dominated by G. holbrooki are likely to be influencing tadpole abundances rather than species richness.

Tail-nipping was observed by G. holbrooki during the single L. olongburensis tadpole experiments. Tail-nipping has been shown to increase the tadpole period of Bombina orientalis (Parichy and Kaplan, 1992) and similar tail-nipping observed by G. holbrooki may result in prolonged periods as a tadpole during the tadpole and metamorph lifestages. Litoria olongburensis (Meyer et al., 2006) and some populations of L. fallax (Anstis, 2002) and Lim. peronii (Simpkins pers. obs.) are associated with ephemeral and semi-permanent waterbodies. Therefore, prolonged time as a tadpole, due to predatory attacks, may be lethal to individual tadpoles of these species as metamorphosis has to occur before waterbody desiccation. This would likely be the same for any ephemeral breeding amphibian species that occurs with G. holbrooki.

Surprisingly, there was no significant difference in predator tadpole consumption between tadpole types for G. holbrooki. There was, however, a significant difference in predator tadpole consumption between tadpole types for Cherax sp. and H. galii. Therefore, the ability of the predators used in this study to influence tadpole populations will differ between tadpole and predator species. Regardless, the introduction of any predator used in this study into a ‘predator’ free waterbody where tadpoles are naïve to the introduced predator may significantly impact tadpole assemblages. This needs to be considered in the construction of permanent waterbodies (i.e. to mitigate against the effects of climate change (Shoo et al., 2011)) when being built for amphibians that lack adequate anti-predator strategies.

Results from the no-choice experiments describe the potential for predation and may only be applicable under specific conditions within the natural environments. Firstly, the presence of alternative prey sources can influence predation rates in experiments (Pyke and 133

White, 2000; Reynolds, 2009) and levels of predation may vary within the natural environment where alternative prey sources would be present. Secondly, aquatic refuge habitat can positively influence survival of tadpoles against predatory fish (Morgan and Buttemer, 1996) and aquatic invertebrate insects (Babbitt and Tanner, 1998; Tarr and Babbitt, 2002; Kopp et al., 2006). Therefore, when present in natural waterbodies, both alternative prey sources and adequate refuge cover can allow for co-existence of anuran species and G. holbrooki. This may be of particular importance to L. olongburensis, that are often associated with waterbodies where sedges are dominant (Simpkins unpublished data). Hence, results from this study may be more applicable to waterbodies where/when refuge habitat and alternative prey sources are relatively low or absent.

Acknowledgements

Work undertaken for this chapter was performed under approval from Griffith University Animal Ethics Committee (Permit number: ENV/18/11/AEC), Queensland general fisheries permit (Permit Number: 90306) and Queensland Department of Environment and Resource Management (ECOACCESS Permit Numbers: WITK10080611 / WISP10081411). I would like to thank Clare Morrison, Donna Treby, James Bone, Diana Virkki, Jean-Marc Hero, Guy Castley and Katrin Lowe for feedback on earlier drafts of this chapter. I also thank Amanda Winzar, Chays Ogston, Jodie Lee Hills, Chris Dahl, Diana Virkki, Tempe Parnell and James Bone for assisting in the field. I also thank Alan Kerr from the Bribie Island Environmental Protection Society provided accommodation during fieldwork. Special thanks are given to Currumbin Wildlife Sanctuary for providing access to their land. Funding was provided from the Griffith School of Environment.

134

5.6 References

Álvarez, D., Nicieza, A. G., 2009. Differential success of prey escaping predators: tadpole vulnerability or predator selection? Copeia 2009, 453-457. Anstis, M. 2002. Tadpoles of South-eastern Australia: a guide with keys, New Holland Publishers, Sydney. Axelsson, E., Nyström, P., Sidenmark, J., Brönmark, C., 1997. Crayfish predation on amphibian eggs and larvae. Amphibia-Reptilia 18, 217-228. Azevedo-Ramos, C., Van Sluys, M., Hero, J.-M., Magnusson, W. E., 1992. Influence of tadpole movement on predation by odonate naiads. Journal of Herpetology 26, 335- 338. Babbitt, K. J., Tanner, G. W., 1998. Effects of cover and predator size on survival and development of Rana utricularia tadpoles. Oecologia 114, 258-262. Baber, M. J. 2001. Understanding anuran community structure in temporary wetlands: the interaction and importance of landscape and biotic processes. Dissertation, Florida International University. Baber, M. J., Babbitt, K. J., 2003. The relative impacts of native and introduced predatory fish on a temporary wetland tadpole assemblage. Oecologia 136, 289-295. Crossland, M. R., 2001. Ability of predatory native Australian ﬁshes to learn to avoid toxic larvae of the introduced toad Bufo marinus. Journal of Fish Biology 59, 319-329. Dorn, N. J., Wojdak, J. M., 2004. The role of omnivorous crayfish in littoral communities. Oecologia 140, 150-159. Furse, J. M., Wild, C. H., 2004. Laboratory moult increment, frequency, and growth in Euastacus sulcatus, the Lamington Spiny Crayfish. Freshwater Crayfish 13, 205- 211. Gillespie, G., Hero, J.-M. (1999) Potential impacts of introduced fish and fish translocations on Australian amphibians. In: Campbell A (ed) Declines and Disappearances of Australian Frogs. Environment Australia, Canberra, pp 131-144 Gregoire, D. R., Gunzburger, M. S., 2008. Effects of predatory fish on survival and behavior of Larval Gopher Frogs (Rana capito) and Southern Leopard Frogs (Rana sphenocephala). Journal of Herpetology 42, 97-103. 135

Gunzburger, M. S., Travis, J., 2005. Critical literature review of the evidence of unpalatability of amphibian eggs and larvae. Journal of Herpetology 39, 547-571. Hamer, A. J., Lane, S. J., Mahony, M. J., 2002. Management of freshwater wetlands for the endangered green and golden bell frog (Litoria aurea): roles of habitat determinants and space. Biological Conservation 106, 413-424. Hecnar, S. J., M'Closkey, R. T., 1997. The effects of predatory fish on amphibian species richness and distribution. Biological Conservation 79, 123-131. Hero, J.-M., Gascon, C., Magnusson, W. E., 1998. Direct and indirect effects of predation on tadpole community structure in the Amazon rainforest. Australian Journal of Ecology 23, 474-482. Hero, J.-M., Magnusson, W. E., Rocha, C. F. D., Catterall, C. P., 2001. Antipredator defenses influence the distribution of amphibian prey species in the central Amazon rain forest. Biotropica 33, 131-141. Heyer, W. R., McDiarmid, R. W., Weigmann, D. L., 1975. Tadpoles, predation and pond habitats in the tropics. Biotropica 7, 100-111. Jara, F. G., 2008. Tadpole–odonate larvae interactions: influence of body size and diel rhythm. Aquatic Ecology 42, 503-509. Kats, L. B., Ferrer, R. P., 2003. Alien predators and amphibian declines: review of two decades of science and the transition to conservation. Diversity and Distributions 2003, 99-110. Kats, L. B., Petranka, J. W., Sih, A., 1988. Antipredator defenses and the persistence of amphibian larvae with fishes. Ecology 69, 1865-1870. Kopp, K., Wachlevski, M., Eterovick, P. C., 2006. Environmental complexity reduces tadpole predation by water bugs. Canadian Journal of Zoology 84, 136-140. Lawler, K. L., Hero, J.-M., 1997. Palatability of Bufo marinus tadpoles to a predatory fish decreases with development. Wildlife Research 24, 327-334. McCollum, S. A., Leimberger, J. D., 1997. Predator-induced morphological changes in an amphibian: predation by dragonflies affects tadpole shape and color. Oecologia 109, 615-621. Meyer, E., Hero, J.-M., Shoo, L., Lewis, B. (2006) National recovery plan for the wallum sedgefrog and other wallum-dependent frog species. Report to the Department of 136

the Environment and Water Resources, Canberra. Queensland Parks and Wildlife Service, Brisbane, Morgan, W. A., Buttemer, L. A., 1996. Predation by the non-native fish Gambusia holbrooki on small Litoria aurea and L. dentata tadpoles. Australian Zoologist 30, 143-149. Nelson, D. W. M., Crossland, M. R., Shine, R., 2011. Foraging responses of predators to novel toxic prey: effects of predator learning and relative prey abundance. Oikos 120, 152-158. Nyström, P. (2002) Ecology. In: Holdich DM (ed) Biology of Freshwater Crayfish. Blackwell Science, Great Britain, pp 192-235. Ohba, S.-y., Nakasuji, F., 2006. Dietary items of predacious aquatic bugs (Nepoidea: Heteroptera) in Japanese wetlands. Limnology 7, 41-43. Parichy, D. M., Kaplan, R. H., 1992. Developmental consequences of tail injury on larvae of the Oriental Fire-Bellied Toad, Bombina orientalis. Copeia 1992, 129-137. Pyke, A. W., White, G. H., 2000. Factors influencing predation on eggs and tadpoles of the endangered Green and Golden Bell Frog Litoria aurea by the introduced Plague Minnow Gambusia holbrooki. Australian Zoologist 31, 496-505. Pyke, G. H., 2008. Plague minnow or mosquito fish? A review of the biology and impacts of introduced Gambusia species. Annual Review of Ecology, Evolution and Systematics 39, 171-191. Renai, B., Gherardi, F., 2004. Predatory efficiency of crayfish: comparison between indigenous and non-indigenous species. Biological Invasions 6, 89-99. Reynolds, S. J., 2009. Impact of the Introduced Poeciliid Gambusia holbrooki on Amphibians in Southwestern Australia. Copeia 2009, 296-302. Saidapur, S. K., Veeranagoudar, D. K., Hiragond, N. C., Shanbhag, B. A., 2009. Mechanism of predator–prey detection and behavioral responses in some anuran tadpoles. Chemoecology 19, 21-28. Shoo, L. P., Olson, D. H., McMenamin, S. K., Murray, K. A., Van Sluys, M., Donnelly, M. A., Stratford, D., Terhivuo, J., Merino-Viteri, A., Herbert, S. M., Bishop, P. J., Corn, P. S., Dovey, L., Griffiths, R. A., Lowe, K., Mahony, M., McCallum, H., Shuker, J. D., Simpkins, C., Skerratt, L. F., Williams, S. E., Hero, J.-M., 2011. 137

Engineering a future for amphibians under climate change. Journal of Applied Ecology 48, 487-492. Skelly, D. K., 1994. Activity level and susceptibility of anuran larvae to predation. Animal Behaviour 47, 465-468. Smith, A. R., Awan, G. R., 2009. The roles of predator identity and group size in the antipredator responses of American toad (Bufo americanus) and bullfrog (Rana catesbeiana) tadpoles. Behaviour 146, 225-243. Stav, G., Kotler, B. P., Blaustein, L., 2007. Direct and indirect effects of dragonfly (Anax imperator) nymphs on green toad (Bufo viridis) tadpoles. Hydrobiologia 579, 85-93. Stoneham, M. 2001. The influence of stream-dwelling predators on the distribution and density of Mixophyes tadpoles in Southeast Queensland. Unpublished Honours Thesis, Griffith University. Tarr, T. L., Babbitt, K. J., 2002. Effects of habitat complexity and predator identity on predation of Rana clamitans larvae. Amphibia-Reptilia 23, 13-20. Touchon, J. C., Warkentin, K. M., 2008. Fish and dragonfly nymph predators induce opposite shifts in color and morphology of tadpoles. Oikos 117, 634-640. Verhoef, G. D., Jones, P. L., Austin, C. M., 1998. A comparison of natural and artificial diets for juveniles of the Australian freshwater crayfish Cherax destructor. Journal of the World Aquaculture Society 29, 243-248. Vonesh, J. R., Kraus, J. M., Rosenberg, J. S., Chase, J. M., 2009. Predator effects on aquatic community assembly: disentangling the roles of habitat selection and post- colonization processes. Oikos 118, 1219-1229. Webb, C., Joss, J., 1997. Does predation by the fish Gambusia holbrooki (Atheriniformes: Poeciliidae) contribute to declining frog populations? Australian Zoologist 30, 316- 324. Webb, C. E. 1994. Does predation by Gambusia holbrooki (Atheriformes: Poeciliidae) contribute to declining frog populations? Unpublished Honours Thesis, Macquarie University. Wells, K. D. 2007. The Ecology and Behavior of Amphibians, The University of Chicago Press, USA.

138

Woodward, B. D., 1983. Predator-prey interactions and breeding-pond use of temporary- pond species in a desert anuran community. Ecology 64, 1549-1555.

139

6.0 General Conclusions

This thesis aimed to determine which environmental variables influenced the adult and tadpole assemblages of ‘acid’ and ‘non-acid’ frog species within and around wallum vegetation of eastern Australia, with primary focus on threatened amphibian species occurring within this habitat. The conclusions are presented from each of four chapters within this thesis, followed by future research directions and management outcomes summarised for the acid frog species, Litoria olongburensis and Crinia tinnula.

6.1 Chapter overviews

6.1.1 Chapter 2 - Variables influencing wallum heathland tadpole assemblages

The tadpole assemblage associated with natural waterbodies in wallum areas included the species L. olongburensis, C. tinnula, L. fallax, L. gracilenta and L. cooloolensis. The assemblage was dominated by tadpoles of L. olongburensis and C. tinnula, indicating that these species are highly adapted to exist within the wallum heathland environment. L. olongburensis were recorded from 11 survey transects while C. tinnula were recorded from 14 survey transects, with both species dominating the waterbodies surveyed. Tadpoles of the other ‘acid’ frog species L. cooloolensis were only recorded from two transects. The absence of this species from other transects are likely to do with its restricted distributional range. The other ‘acid’ frog species L. freycineti was absent from the tadpole assemblage. This is likely because of detectability issues with this species.

The distribution and occupancy of L. olongburensis and C. tinnula tadpoles within natural wallum heathland waterbodies was associated with several environmental variables, with variables differing between species, distribution and occupancy. The variables influencing L. olongburensis abundance included pH, depth and turbidity while the variables influencing L. olongburensis occupancy included pH and depth. Both pH and depth were 140

influencing L. olongburensis tadpole in a unimodel distribution. The variables influencing C. tinnula abundance were not definitive while the primary variables influencing C. tinnula occupancy included positive associations with water depth, and turbidity, and a negative association with predatory fish.

These results confirm that wallum heathland waterbodies are dominated by ‘acid’ frog species (L. olongburensis and C. tinnula). Furthermore, it confirms hypothesis by Meyer et al (2006) that indicated that pH is playing a role in L. olongburensis distribution. The lower relative abundance of L. fallax is likely due to tadpole intolerance to low pH waters (Meyer 2004).

6.1.2 Chapter 3 - Usage of anthropogenic waterbodies, and variables influencing adult amphibian assemblages

Natural waterbodies contained L. olongburensis, L. fallax, L. tyleri, C. tinnula, L. gracilenta and Limno. peronii while anthropogenic waterbodies contained L. olongburensis, L. fallax, L. tyleri, C. tinnula, L. gracilenta, R. marina, L. nasuta, L. freycineti, a Uperolia sp. and Limno. peronii. The waterbody type (i.e. golf course, road side ditch) that a species was located in was species specific.

The relative abundance of L. olongburensis within natural and anthropogenic waterbodies was associated with both both % sedge cover and pH. The result of pH is similar for L. olongburensis tadpole distribution within Chapter 2 and adult relative abundance in Chapter 4, indicating that water pH is vital for L. olongburensis population continuance in wallum heathland waterbodies. Furthermore, L. olongburensis and C. tinnula relative abundance and L. olongburensis tadpole occupancy was highest in natural waterbodies.

Results from this chapter also showed that an overlap in the ecological niches for L. olongburensis and L. fallax occur. This indicates co-existence between these two species within certain waterbodies. However, high pH had a negative influence on L. olongburensis

141

abundance, indicating that tadpole competition between L. olongburensis and L. fallax may be excluding L. olongburensis within waterbodies where pH is high. Additionally, relative abundance of L. fallax, while not shown statistically, was lower in waterbodies where pH was low, suggesting L. fallax tadpoles are either intolerant to low pH or are outcompeted by L. olongburensis in acidic waters. The intolerance hypothesis is likely correct, as Meyer (2004) showed eggs and tadpoles of L. fallax fail to metamorphose when exposed to low pH water (i.e. 3.5.). Furthermore, the absence of adults from ‘unfavourable’ waterbodies could be explained by adults failing to deposit eggs in waterbodies that are unfavourable for successful reproduction (Takahashi, 2007; Hamamura, 2008).

6.1.3 Chapter 4 – Compensatory pond usage by wallum heathland amphibians and variables influencing adult amphibian assemblages

In a similar study of natural and compensatory waterbodies of a four-lane road construction project, environmental variables associated with the relative abundance of L. olongburensis were pH, salinity and water depth (similar to the distribution of L. olongburensis tadpoles in natural wallum habitats (Chapter 2) and disturbed waterbodies (Chapter 3). The result of water depth is similar for L. olongburensis tadpole distribution within Chapter 2. This gives a strong indication that both of these variables are important for L. olongburensis population persistence.

Environmental variables associated with the relative abundance of C. tinnula were salinity, water depth, minimum hydroperiod and pH. The only variable influencing C. tinnula tadpole occupancy (Chapter 2) that also influenced the relative abundance of C. tinnula adults was water depth, indicating that this is an important variable in C. tinnula population persistence.

The lower relative abundance of ‘acid’ frog species in compensatory habitats is explained by unfavourable water chemistry variables and the high relative abundance of competitive amphibian species. These results support the conclusions of Chapter 3, suggesting

142

competition excludes the ‘acid’ frogs from environments where water chemistry is more favourable for competitive amphibian species.

6.1.4 Chapter 5 - Predation experiments with G. holbrooki and natural predators

Predation experiments suggest G. holbrooki are either less efficient or equally efficient at predating on tadpoles when compared with native predatory species. This was dependant on tadpole species. For L. olongburensis, the most effective predators were Cherax sp. and Belostomatida, with G. holbrooki predation being low or absent. This trend was similar for the other two tadpole species (L. fallax and Limno. peronii), with G. holbrooki predation being low or equal to predation rates by native predatory species.

While these results indicate that G. holbrooki have a lower predation potential when compared with native predaotrs, itshould be noted that, when present, G. holbrooki are often the most abundant fish species (reviewed in Pyke, 2008) and would therefore increase the occurrence of predator-prey interactions and potentially influence the abundance of amphibians at the waterbody. However, amphibian species richness is unlikely to be influenced by presence of G. holobrooki, with past field studies finding that presence of L. aurea tadpoles (Hamer et al., 2002) and adults of other Australian native amphibian species (Reynolds, 2009) were not influenced by the presence of G. holbrooki.

Results from previous chapters also showed that G. holbrooki presence within natural wallum waterbodies were low in occupancy (Chapter 2), indicating successful colonization of G. holbrooki in remote, natural waterbodies is low.

6.2 Management Outcomes

Results from this thesis showed that generalist amphibian species dominate the amphibian assemblages occurring in anthropogenically disturbed areas surrounding wallum heathland

143

areas (Chapter 3 + 4). The opposite was found in natural waterbodies within wallum heathland, where a few specialist ‘acid’ frog species dominating the amphibian assemblage (Chapter 2). For L. olongburensis and C. tinnula, low tadpole occupancy within anthropogenic waterbodies and high occupancy/relative abundance within natural waterbodies indicates that ‘acid’ frogs are predominantly restricted to natural waterbodies for breeding. Therefore, when habitat offsets are considered for L. olongburensis and/or C. tinnula, natural waterbodies should be targeted for conservation. Furthermore, compensatory waterbodies should exclude or have low relative abundance of native predatory fish and ensure optimal water chemistry and vegetation characteristics. These characteristics should be within the ranges where L. olongburensis and C. tinnula adults and tadpoles were predominantly encountered during this survey (Chapters 2-4).

The first two years after construction of compensatory ponds are inadequate in providing habitat for the acid frog species L. olongburensis and C. tinnula. However, compensatory habitats will provide habitat for non-target competitive species (i.e. L. fallax). As mentioned previously, failure of compensatory habitats providing adequate habitat for ‘acid’ frogs is probably a result of unfavourable environmental conditions. If future construction of compensatory habitats is undertaken then managers should target areas where environmental variables (i.e. pH) would be within the ideal ranges for acid frog species.

While field surveys indicated that L. olongburensis tadpoles prefer waterbodies where G. holbrooki are absent, no-choice experiments suggest that native predatory fauna have the ability to predate on tadpoles of L. olongburensis more than G. holbrooki. Despite this, G. holbrooki are often high in abundance in waterbodies where L. olongburensis are present (Pyke, 2008). Hence, G. holbrooki may influence the abundance, rather than the occupancy, of L. olongburensis. The interactions between G. holbrooki and other predators in these ecosystems needs to be further examined. Where possible, there should be no introduction of predatory fish or aquatic tadpole predators into newly constructed waterbodies or natural wallum heahtland waterbodies as fish are likely to be negatively influencing the abundance of L. olongburensis tadpoles. 144

6.3 Future priorities for research

Field studies conducted for this thesis were in the summer/autumn period. Hines and Meyer (2011) recorded over 20 species of amphibians occurring on south-east Queensland dune islands. These islands contain both wallum and other native vegetation communities. Future studies need to be conducted during different seasons to identify winter breeding and potentially missed amphibian species that were not detected during these surveys. This should be undertaken in both natural and anthropogenic waterbodies to capture a ‘disturbance’ gradient to determine if any missed species are utilizing anthropogenic waterbodies for breeding.

Additional surveys determining what environmental variables influence the occupancy and abundance of other acid frog species (i.e. L. freycineti and L. cooloolensis) are needed. This would not just be targeted to mainland populations, but be extended to the dune islands what are located within south-east Queensland.

Prolonged field surveys targeting waterbodies that were designed to aid in compensatory habitats for ‘acid’ frogs are required to determine if compensatory habitats can be successfully utilized by ‘acid’ frogs after a two year period. Ecological studies are often confined to relatively short survey periods and these may not provide an accurate assessment of the long-term patterns or trends within populations. Therefore future periodic monitoring is required to address research questions related to the long-term survival of these species. Furthermore, studies should be undertaken at differing latitudes and environments across the distributional range of both L. olongburensis and C. tinnula.

A thorough examination of the predator – prey interactions among all predators and prey to determine how fish and other predators interact, and clarify the relationships between acid frogs and fish, are required. Alternative prey sources and aquatic refuge cover can influence predation rates on tadpoles by aquatic predators. Therefore, predation experiments using alternative prey and a varying density of tadpoles and predators need to be conducted. This

145

would allow for results to be further extrapolated to all waterbodies within wallum heathland environments and not limited to waterbodies where refuge cover and alternative prey sources are low or absent.

6.4 References

Hamer, A. J., Lane, S. J., Mahony, M. J., 2002. Management of freshwater wetlands for the endangered green and golden bell frog (Litoria aurea): roles of habitat determinants and space. Biological Conservation 106, 413-424. Haramura, T., 2008. Experimental test of spawning site selection by Buergeria japonica (Anura: Rhacophoridae) in response to salinity levels. Copeia 2008, 64-67. Hines, H.A., Meyer, E.A., 2011. The frog fauna of Bribie Island: an annotated list and comparision with other Queensland dune islands. Proceedings of the Royal Society of Queensland 117, 261-274. Meyer, E., 2004. Acid adaptation and mechanisms for softwater acid tolerance in larvae of anuran species native to the “Wallum” of east Australia. PhD Thesis, University of Queensland. Pyke, G. H., 2008. Plague minnow or mosquito fish? A review of the biology and impacts of introduced Gambusia species. Annual Review of Ecology, Evolution and Systematics 39, 171-191. Reynolds, S. J., 2009. Impact of the Introduced Poeciliid Gambusia holbrooki on Amphibians in Southwestern Australia. Copeia 2009, 296-302 Takahashi, M., 2007. Oviposition site selection: pesticide avoidance by gray treefrogs. Environmental Toxicology and Chemistry 26, 1476-1480.

146

7.0 Appendices: Publications on ‘acid’ frogs published during candidature

Appendix 1

Simpkins, C.A., Meyer, E., Hero, J.-M., 2011. Long-range movement in the rare Cooloola sedgefrog Litoria cooloolensis. Australian Zoologist 35:977-978.

Understanding habitat usage is essential for the proper management of rare and threatened species in the wild. However, current knowledge of habitat usage by many rare and threatened Australian frog species is inadequate in this regard. Knowledge of non-breeding habitat usage in Australian amphibian species is particularly poor (Hines et al. 1999), with current understanding of the habitat requirements of many species based largely on habitat usage by calling animals during the breeding season, when frogs are more readily detectable. However, like other fauna, the use of different habitats by amphibians can vary daily and seasonally as well as between different sexes and life stages (Law and Dickman 1998). Female frogs, for example, are likely to occupy habitat further away from a breeding water body while males tend to stay near the breeding site (Bartelt et al. 2004; Johnson et al. 2007; Rittenhouse and Semlitsch 2007). A majority of aquatic breeding amphibians will also undergo embryonic and larval development within the aquatic environment while spending their adult and juvenile lives in the terrestrial environment, unlike developing larvae, post-metamorphic adult and juvenile amphibians are not restricted to the aquatic body of their birth (Johnson et al 2007) and thus have the ability to disperse or migrate into adjacent, non-breeding, areas for the purposes of foraging, overwintering (Regosin et al 2003) or refuge use (Semlitsch and Bodie 2003).

The Cooloola sedgefrog (Litoria cooloolensis) is a rare, largely arboreal species restricted to coastal wallum (i.e., coastal sand dunes and plains) in south-east Queensland. Though known to occur 'some distance' away from breeding habitat within the terrestrial landscape

147

(Meyer et al. 2006), details of non-breeding habitat usage in this species (e.g., the type of habitat utilized by non-breeding animals and movement distances from areas of breeding habitat) remain poorly documented. Herein, we present detailed observational records of L. cooloolensis from non-breeding areas in wallum habitat, thus providing additional insight into non-breeding habitat usage by this species.

On a summer's night in late January-early February 1999, upwards of 50 L. cooloolensis were heard calling from trees along a walking track through rainforest near Lake Poona, within the Cooloola section of the Great Sandy Region National Park, south east Queensland. The majority of animals heard calling at this time appeared to be calling high up in the forest canopy, producing the same loud 'kik' call as that made by male L. cooloolensis heard calling later at Lake Poona and elsewhere (EM, pers. obs.; Naturesound, 1998). While most animals were heard calling from forest habitat close to Lake Poona (i.e., within 100 metres of suitable breeding habitat), between 30-50 animals were heard calling much further (> 300 m) from water. This includes 5-10 animals heard calling from rainforest at Bymien Picnic Area (-25.95447 153.1045°) approximately 900 m from Lake Poona, the nearest known breeding site for this species. On the same night, numerous L. cooloolensis (upwards of 200 animals) were heard and/or seen calling from vegetation fringing at Lake Poona (-25.9637° / 153.1104°). Male calling activity within forest habitat surrounding Lake Poona, including rainforest at Bymien Picnic Area, continued well past midnight.

Small numbers of L. cooloolensis (between 10 and 20 animals) were again heard calling from rainforest along the same walking track to Lake Poona on the evening of January 29, 2000. As before, animals were heard calling from Bymien Picnic Area at the start of the walking track to Lake Poona, 900 m from the lake itself Large numbers of L. cooloolensis (upwards of 200 animals) were again heard and/or seen calling at Lake Poona. Male calling activity within forest habitat away from Lake Poona (as far away as Bymien Picnic Area) continued through to dawn. Diurnal calling of L. cooloolensis was also opportunistically recorded in the southern Cooloola section of the Great Sandy Region National Park (- 26.2389° / 153.0684°) on the 29th April, 9th July and the 13th July of 2010 in open 148

Eucalyptus forest trees at an elevation of approximately 95 metres above sea level. The calls made by animals at this time were the same as described above. In contrast with the above observations, calling activity was short-lived, lasting no longer then 30 seconds. The maximum number of calling individuals at this site was two (recorded on the 29th April 2010). The nearest known breeding habitat for L. cooloolensis is a wetland/coastal swamp located approximately 1.3 kilometres from the location of the calling individuals (-26.2351° / 153.0556°). No other suitable breeding habitat was located near this site during surveys conducted along a 5km east-west transect or on examination of Google Earth satellite images.

The presence of L. cooloolensis up to 1.3 kilometres from suitable breeding habitat demonstrates this species is capable of moving widely across forested terrain. The large numbers of animals heard calling from rainforest near Lake Poona, moreover, suggests forest surrounding breeding areas can, at times, support large numbers of this rare species. The loss or degradation of forest habitat in wallum areas occupied by L. cooloolensis (e.g., as a result of intense wildfire) may therefore have a significant impact on numbers of this species and, potentially, the ability of animals to move within the broader landscape.

The ability of L. cooloolensis to move large distances is a trait shared with the closely- related common sedgefrog (Litoria fallax), a species often encountered in dry eucalypt forest hundreds of metres from suitable breeding habitat (CAS, EM, J-MH pers. obs). Like L. cooloolensis, L. fallax can often be heard calling from the canopy of forest trees, occasionally on dry ridges many hundreds of metres from water (EM pers. obs.). Why these small, diminutive (SVL <45 mm) frogs should travel so widely is presently unclear. In the case of L. cooloolensis, movement of animals away from breeding areas could occur in response to: (1) increased competition amongst males at breeding sites during times of heightened calling activity; or (2) limited availability of food in breeding areas (due to increased competition with congeners at breeding sites during peaks in breeding activity). As well as helping animals avoid competition, large scale movements of L. cooloolensis could potentially promote gene flow between isolated populations, as well as facilitate 149

establishment of new breeding populations (Krebs 2009). Further data on competition, movement and gene flow in L. cooloolensis are needed to determine which, if any, of these scenarios is correct.

Though male frogs may call for a number of reasons, most call either to attract a mate (i.e., for breeding purposes) or to advertise their presence to other males (i.e., for territorial purposes) (Zug et al. 2001; Wells 2007). In the case of L. cooloolensis, calling at distance from breeding sites is unlikely to facilitate breeding, as females mating with males remote from areas of breeding habitat would have to move many hundreds of metres in order to spawn. Observations of more than one calling L. cooloolensis in the described areas suggests observed calling activity away from breeding sites could instead be a territorial response towards other male L. cooloolensis in the area. Whether this calling behaviour represents practiced territorial calling in response to calling by conspecific males or calling for some other purpose (e.g., defence or delineation of foraging territories) (see e.g., Ryan 2001) is presently unclear as visual observations of calling individuals, remote from breeding areas, were not made. More detailed study of the calling and social behaviour of this species are needed to ascertain the functional significance of calling by animals in non-breeding habitat.

References

Bartelt, P.E., Peterson, C.R. and Klaver, R.W, 2004. Sexual differences in tbe post- breeding movements and habitats selected by westem toads {Bufo bóreas) in southeastern Idaho. Herpetologica 60(4): 455-467.

Hines, H.B., Mahony, M. and McDonald, K., 1999. An assessment of frog declines in wet subtropical Australia. Pp 44-63 in Campbell, A. (ed.) Declines and Disappearances of Australian Frogs. Environment Australia, Canberra.

Johnson, J.R., Knouft, J.H. and Semlitsch, R.D., 2007. Sex and seasonal differences in tbe spatial terrestrial distribution of gray treefrog (Hyla versicolor) populations. Biological Conservation 140: 250-258.

150

Krebs, C.J., 2009. Ecology: The Experimental Analysis of Distribution and Abundance: Intemal Edition. Pearson Education, United States of America.

Law, B.S. and Dickman, C.R., 1998. The use of habitat mosaics by terrestrial vertebrate fauna: implications for conservation and management. Biodiversity and Conservation 7: 323-333.

Meyer, E., Hero, J.-M., Shoo, L. and Lewis, B., 2006. Recovery plan for tbe wallum sedgefrog and otber wallum- dependent frog species 2005-2009. Report to Department of Environment and Heritage, Canberra. Queensland Parks and Wildlife Service, Brisbane.

Regosin, J. V, Windmiller, B.S. and Reed, J.M., 2003. Terrestrial Habitat Use and Winter Densities of tbe Wood Frog (Rana sylvatica). Joumal of Herpetology 37(2): 390-394.

Rittenhouse, T.A.G. and Semlitsch, R.D., 2007. Distribution of amphibians in terrestrial habitat surrounding wetlands. Wetlands 27(1): 153-161.

Ryan, M., 2001. Anuran Communication. Smithsonian Institute Press, Washington.

Semlitsch, R.D. and Bodie, J.R., 2003. Biological Criteria for Buffer Zones around Wetlands and Riparian Habitats for Amphibians and Reptiles. Conservation Biology 17(5): 1219-1228.

Wells, K.D., 2007. The Ecology and Behavior of Amphibians. The University of Chicago Press, Chicago.

Zug, G.R., Vitt, L.J. and Caldwell, J.P, 2001. Herpetology: An Introductory Biology of Amphibians and Reptiles, Second Edition. Academic Press, San Diego.

151