Heleioporus Albopunctatus) in the Fragmented Landscape of The

Metapopulation structure of the Western Spotted Frog

(Heleioporus albopunctatus) in the fragmented landscape of the

Western Australian wheatbelt.

Robert A. Davis BSc (Hons)

Department of Zoology

The University of Western Australia

This thesis is presented for the degree of Doctor of Philosophy of The

University of Western Australia

2004

Foreword

This thesis has been written as a series of self-contained scientific papers, the majority of which are currently being considered for publication in peer-reviewed journals. At the time of writing two co-authored papers (chapters 2 and 3) have been provisionally accepted for publication, pending minor amendment. I contributed 95% to both chapters whilst the second author contributed 5%. To the remaining chapters, I contributed 100%. Publications are as follows:

Davis, R.A. and Roberts, J.D. The population genetic structure of the Western Spotted

Frog, Heleioporus albopunctatus (Anura: Myobatrachidae) in a fragmented landscape in south-western Australia. Australian Journal of Zoology. (accepted pending minor amendments).

Davis, R.A. and Roberts, J.D. Embryonic survival and egg numbers in small and large populations of the frog Heleioporus albopunctatus in Western Australia. Journal of

Herpetology. (in press, 2005).

Additionally, another published paper arising from (but not a direct part of) this PhD study is appended (Appendix 1).

The order of this thesis is such that each paper naturally leads to the other, and all chapters (papers) refer frequently to each other in the text. I have prepared a general introduction outlining the context of this study. Some repetitive introductory material from each chapter has been removed and placed in the general introduction. All references are placed at the end of the thesis.

Acknowledgements

This project would not have been possible without the support of a great number of individuals and organisations. Above all I would like to thank my supervisor Dale

Roberts for his friendship, guidance, constructive criticism, field assistance and insight. Dale reviewed all chapters of this thesis and they have greatly benefited from his input. My partner Jennifer Wilcox has put up with this project for longer than anyone should have to and I thank her for her love, wisdom, inspiration and friendship amongst the chaos. My long-suffering roommate Wes Bancroft has been a continual source of inspiration and I thank him for his friendship and for brightening my day.

Primary funding for this research was provided by the Department of Zoology,

UWA. I thank the following funding bodies for providing further financial assistance: the Peter Rankin Trust for Herpetology (The Australian Museum), the Australian

Geographic Society and The Butler Trust of the Museum of Western Australia. The

Department of Conservation and Land Management provided fauna licences

NE002947 and SF003756. The UWA Animal Ethics Committee approved all work under licence 99/008/E92.

A number of individuals have helped me with statistical advice, laboratory techniques, figures and other questions. I am indebted to Mike Bamford, Wes

Bancroft, Christine Cooper and Alex Larcombe for assistance with figure preparation.

Bob Black, Jane Prince and Phil Withers assisted with queries on experimental design and statistics. Without the assistance of Mike Johnson, Terrie Finston and Michelle

Stuckey I would have been unable to conquer the allozyme lab! I thank them profusely. Bert and Barbara Main from UWA imparted their legendary frog knowledge to me on many occasions. Ken Aplin, Mark Cowan, Ric How and Brad

Maryan of Terrestrial Vertebrates at the Museum of Western Australia provided information, advice and access to museum collections. Richard Hobbs of Murdoch

University and Michael and Lesley Brooker of CSIRO Sustainable Ecosystems provided information and advice. Rick Roberts, Cameron Duggin and Hai Ngo from

UWA Zoology provided cheerful friendship and technical assistance.

Colleagues that kindly assisted with reviewing drafts of manuscripts included Paul

Doughty, Don Driscoll, Martin Dziminski and Josh Van Buskirk. I thank Henry

Disney of Cambridge University, England, for identifying dipteran specimens and co- authoring our paper.

I am indebted to a number of friends for providing their advice, friendship and physical assistance with fieldwork. Catherine Arrese, Wes Bancroft, Martin

Dziminski, Carole Elliott, Brad Maryan, Brenden Metcalf and Jennifer Wilcox all brightened my fieldwork with their enthusiastic assistance. The transient UWA Herp

Lab provided a continual source of coffee, friendship and comradeship. I especially thank Mike Smith and Phil Byrne for showing me how to live life!

This project would have been impossible without the co-operation and generosity of the landowners and people of Kellerberrin who offered me accommodation, friendship and assistance when I needed it most. I am particularly indebted to Colin

Wilkins and family, Frank and Noelene Morley, Rod and Judy Forsyth and family,

Kit and Eileen Leake and family and Gavin and Amanda Morgan and family. Tom

Groves also kindly permitted me to work on his land.

I thank my family Brian, Ros and Annabelle Davis for always being there for me, and finally to Lexie for always getting in my way when I was trying to type.

Abstract

Amidst concern over the global phenomenon of declining amphibians, there is an increasing appreciation of the importance of understanding population dynamics at both local and regional scales. Data on the viability and persistence of species in landscapes altered by humans are scarce but an understanding of these dynamics is essential for enabling long-term species conservation in a modified world. Habitat loss, fragmentation and ensuing salinisation are of particular concern for species in

Australia’s temperate agricultural regions where the rapid conversion of continuously vegetated landscapes to small fragments has occurred in less than 200 years. This thesis investigated the local and metapopulation structure of Heleioporus albopunctatus to determine the current population structure and likely future of this species in a highly degraded landscape: the wheat and sheep growing areas of south- western Australia.

To investigate gene flow and population subdivision, I examined genetic variation at four variable loci in twenty-two populations using cellulose acetate electrophoresis

(Chapter 2). I found a moderate, but significant degree of subdivision (Fst = 0.087 ±

0.049, p<0.05) across all populations, and high levels of heterozygosity (H = 0.133,

SE = 0.084). Several populations had higher Fst values in pair-wise comparisons. A significant but weak relationship was found between genetic distance and geographic

2 distance (r = 0.019, F1,229 = 4.32, p = 0.039) but this combined with data from multidimensional scaling analyses, revealed that geographic isolation of populations is not a significant determinant of genetic structuring. The regional persistence of this species may be dependent on the maintenance of current or historical metapopulation structures that allow gene flow.

In Chapter 3, I examined the hypothesis that small populations in marginal habitats arising from fragmentation and increased salinity, have higher embryonic mortality and lower clutch sizes than larger populations. Although four of 55 clutches examined

(~ 7%) were infested by dipteran larvae, embryonic mortality was low, averaging 3% per clutch. I found no significant relationship between clutch size or embryonic mortality and population size.

Success at the population level is determined primarily by the number and quality of metamorphosing larvae successfully recruited into the adult population. I investigated the recruitment success of populations of H. albopunctatus breeding in a range of ephemeral pools all derived from human modification of the landscape: e.g salinity interceptor banks (Chapter 4). The three-year recruitment success of 48 ponds monitored was poor with on average only 13.89% of ponds producing metamorphs.

Hydroperiod was the most important determinant of recruitment success indicating population regulation at a local scale, with regional implications. Continued recruitment failures will have a profound impact on the survival of local H. albopunctatus populations, potentially resulting in local extinction as populations age or suffer inbreeding effects.

I conducted a three year mark-recapture study of five breeding populations to quantify the variance in demographic parameters of a non-declining frog species.

Adult survival ranged from 0.34 to 1. Recapture rates were low and ranged from 0.05 to 0.45. Sex ratios and estimated population sizes fluctuated greatly between years and a strong relationship existed between rainfall during the reproductive season and population size (r2 = 0.53 – 1.00).

From a management perspective it is critical to be able to predict the long-term survival of this species in a human-altered landscape. In Chapter 6, I used life-history

data (Chapters 2-5) to undertake a Population Viability Analysis of H. albopunctatus populations to establish how local populations and metapopulations responded to variations in demographic parameters and climatic events such as drought. Juvenile survival was critical to the long-term persistence of H. albopunctatus at a regional and local scale (see also Chapter 4) but the formation of a metapopulation was sufficient to buffer H. albopunctatus against extinction in the long-term. The life-history attributes of H. albopunctatus, including high fecundity, high adult longevity and low to moderate dispersal contribute to a robust regional metapopulation, responsive to changes, but with a strong chance of persistence over the long-term.

H. albopunctatus appears to have adjusted to a radically modified landscape but its long-term persistence may be dependent on the existence of a small number of source populations that recruit in most years.

Foreword...... 1 Acknowledgements ...... 2 Abstract...... 4

List of Tables ...... 10

List of Figures...... 12

Chapter 1: General Introduction ...... 14

STUDY SPECIES ...... 16 STUDY OBJECTIVES ...... 17 Chapter 2: The population genetic structure of the Western Spotted Frog, Heleioporus albopunctatus (Anura: Myobatrachidae) in a fragmented landscape in south-western Australia...... 19

INTRODUCTION ...... 19 MATERIALS AND METHODS ...... 20 Sampling methodology...... 20 Laboratory methodology...... 21 Analyses...... 22

RESULTS ...... 23 Genetic structuring...... 23 Multidimensional Scaling...... 24 Geographic Separation...... 24 Drainage Patterns...... 25

DISCUSSION ...... 26 Metapopulation models for H. albopunctatus...... 30

Chapter 3: Embryonic Survival and Egg Numbers in Small and Large Populations of the Frog Heleioporus albopunctatus in Western Australia ...... 39

INTRODUCTION ...... 39 METHODS...... 41 Egg Survival and Population Size ...... 41 Egg Survival and Developmental Stage ...... 42 Clutch Size and Population Size...... 42 Female Body Size and Clutch Size...... 42

RESULTS ...... 43 Egg Survival and Population Size ...... 43

Egg Survival and Developmental Stage ...... 43 Clutch Size and Population Size...... 44 Female Body Size and Clutch Size...... 44

DISCUSSION ...... 44 Clutch Size and Population Size...... 44 Female Body Size and Clutch Size...... 45 Egg Survival and Population Size ...... 45 Egg Survival and Developmental Stage ...... 46 Causes of Egg Mortality ...... 47

Chapter 4: Larval recruitment success of the frog Heleioporus albopunctatus in a West Australian agricultural landscape...... 53

INTRODUCTION ...... 53 METHODS...... 55 Study Area and Site Selection ...... 55 Recruitment to Tadpoles...... 55 Tadpole Abundance ...... 56 Recruitment to metamorphosis ...... 56 Statistical analyses ...... 57

RESULTS ...... 57 Distribution of anthropogenic ponds ...... 57 Recruitment to tadpoles ...... 58 Tadpole abundance ...... 58 Recruitment to metamorphosis ...... 59 Statistical analyses ...... 59

DISCUSSION ...... 60 Recruitment Success to metamorphosis ...... 60 Hydroperiod and Pond Features ...... 62 Impacts of salinity...... 63 Metapopulation persistence and management implications ...... 64

Chapter 5: Population ecology of the frog Heleioporus albopunctatus in the central wheatbelt of Western Australia: Survival and population size of adults.82

INTRODUCTION ...... 82 METHODS...... 83 Selection of Study Sites ...... 83 Trapping Methods...... 85

Marking and Sexing...... 85 Survival Rate Estimation ...... 85 Population Size Estimation...... 87 Sex Ratios ...... 88

RESULTS ...... 88 DISCUSSION ...... 90 Survival and recapture estimates ...... 90 Population Size ...... 92

Chapter 6: Persistence and extinction risk of the Western Spotted Frog Heleioporus albopunctatus in a fragmented agricultural landscape: a population viability analysis...... 103

INTRODUCTION ...... 103 METHODS...... 104 Program Used for PVA...... 104 Estimation of life-history traits...... 104 Constructing Basic Models...... 106 Scenario Projections ...... 108 Statistical Analysis...... 109

RESULTS ...... 109 Basic Models...... 109 Scenario Projections ...... 110

DISCUSSION ...... 111 Chapter 7: Overview...... 118

References...... 122

List of Tables

Chapter 2...... 19

Table 2.1 ...... 31 Locality and size of breeding populations sampled for genetics Table 2.2 ...... 33 Allele frerquencies for 22 populations of H. albopunctatus for the variable loci detected Table 2.3a...... 35

Pairwise Fst estimates for all population pairs. Table 2.3b ...... 36 Corresponding pairwise variance for 2.3a

Chapter 3 ...... 39

Table 3.1 ...... 49 Reported embryonic mortality in the anuran literature classified according to egg deposition mode

Chapter 4...... 53

Table 4.1 ...... 76 Recrutment success of H. albopunctatus for all ponds containing water, 2000- 2002 Table 4.2 ...... 78 Hydroperiods of 48 ponds monitored from 2000-2002 Table 4.3 ...... 79 Correaltion matrix of pond parameters of 19 breeding ponds monitored in 2000 Table 4.4 ...... 80 Results of PCA of pond parameters affecting metamorphosis including biotic and abiotic variables. Table 4.5 ...... 80 Results of PCA of pond parameters affecting metamorphosis including abiotic variables.

Chapter 5 ...... 82

Table 5.1 ...... 96

The most parsimonious models for four populations of H. albopunctatus Table 5.2 ...... 97 Annual census population size, Jolly-Seber population estimates and sex ratios Table 5.3 ...... 98 Documented annual survival rate ranges for anurans

Chapter 6...... 103

Table 6.1 ...... 115 Basic model input parameters for Vortex simulations of H. albopunctatus population viability Table 6.2 ...... 116 Summary of input parameters for PVA Table 6.3 ...... 117 Results of PVA simulation scenarios for H. albopunctatus

List of Figures

Chapter 2...... 19

Figure 2.1 ...... 37 Location of H. albopunctatus genetic sampling sites in south-west WA and the central wheatbelt Figure 2.2 ...... 38

Multidimensional scaled plot of pairwise Fst differences between populations

Chapter 3 ...... 39

Figure 3.1 ...... 51 The relationship between the number of dead eggs per clutch and population size Figure 3.2 ...... 52 Percentage embryonic mortality in relation to Gosner stage

Chapter 4...... 53

Figure 4.1 ...... 68 Location of northern and southern tadpole study sites north of Kellerberrin, WA, in relation to native vegetation remnants Figure 4.2 ...... 69 Maximum number of tadpoles present at each site sampled from 2000-2002 Figure 4.3 ...... 70 Number of tadpoles present on each sampling occasion, at each of 48 ponds monitored from 2000-2002 Figure 4.4 ...... 71 Relationship between hydroperiod and tadpole abundance for 2000 and 2001 Figure 4.5 ...... 72 Relationship between water conductivity and tadpole abundance for all ponds in which H. albopunctatus tadpoles were present 2000 and 2001 Figure 4.6 ...... 73 Number of metamorphs recruited at 3 ponds monitored for a four year period Figure 4.7 ...... 74

PCA plot of PC1 vs PC2 for full model of variables influencing recruitment success of ponds monitored in 2000. Figure 4.8 ...... 75 PCA plot of PC1 vs PC2 for abiotic model of variables influencing recruitment success of ponds monitored in 2000.

Chapter 5 ...... 82

Figure 5.1 ...... 100 Location of Heleioporus albopunctatus population study sites and major draingage systems, near Kellerberrin in Western Australia. Figure 5.2 ...... 101 Survival and recapture estimates for the three populations with constant survival and recapture phi(.)p(.) Figure 5.3 ...... 102 Survival and recapture estimates for Morley West, based on model averaging of the most parsimonious models phi(g)p(.) and phi(g)p(t)

Chapter 1: General Introduction

Our contemporary natural world is a dramatically different place from that of our ancestors. Every natural ecosystem has been altered by humanity, many to the point of collapse (Meffe and Carroll 1994). It is now estimated that extinction rates are up to 10,000 times the expected background rate (Pimm et al. 1995). Of particular concern amidst the cloud of extinctions, is the recent global phenomenon of catastrophic amphibian declines in populations from diverse geographical localities, including remote wilderness areas on most continents (eg. Barinaga 1990; Blaustein et al. 1994).

It has now been recognized that amphibians are part of a distinctive decline phenomenon that goes beyond the general biodiversity crisis (Pechmann and Wilbur

1994). Whilst much attention has focused on global causes such as climate change

(Pounds and Crump 1994) and disease (e.g. Laurance et al. 1996 ), recent studies have recognised the significance of threats to local populations and the importance of local population dynamics in regional persistence (ie. metapopulation dynamics) (Wake

1998; Alford and Richards 1999; Semlitsch 2002).

Investigations into amphibian declines have revealed the lack of data on basic population parameters (e.g. Pechmann and Wilbur 1994; Alford & Richards 1999).

Detailed demographic studies of amphibians are scarce and fewer are for long time- periods or focused on more than one population (but see Ramirez et al. 1998; Conroy

2001; Gillespie 2001; Marunouchi et al. 2002; Conroy and Brook 2003). Alford and

Richards (1999) have highlighted the need for further intensive studies to obtain basic autecological data on amphibians and the importance of considering populations as a network of local populations (ie. a metapopulation) rather than a single entity. Alford and Richards (1999) remarked that since amphibians often live in metapopulations,

local population declines and extinctions may be common events. In a regional context these events may not be an issue if populations are connected by dispersal. In fragmented landscapes, e.g. generated by clearing for agriculture, dispersal may be interrupted. Any understanding of how populations are to persist at a regional scale, therefore requires an understanding of current metapopulation biology i.e. the opportunities for recruitment and dispersal.

Many amphibian populations have been studied and described in the context of metapopulation theory (eg. Berven and Grudzien 1990; Sjogren 1991, 1994; Driscoll et al. 1994; Driscoll 1997; Driscoll 1998). It is now recognized that metapopulations are critical to the survival of many amphibian species and those species with aspects of biology or behaviour that are not conducive to dispersal, may inevitably succumb to extinction (eg. Driscoll 1997).

Of the many factors affecting amphibian populations, the impacts of habitat fragmentation have received little attention. This may be due to the less obvious, long-term impacts of fragmentation such as restricted gene flow and subsequent erosion of heterozygosity, rather than the catastrophic impacts of disease and introduced predators. Habitat fragmentation is of particular concern for amphibians whose range encompasses extensively human-modified landscapes including agricultural regions and urban areas. A number of studies have found that amphibians are sensitive to habitat fragmentation (Laan and Verboom 1990; Wyman 1990;

Sjogren 1991; Wake 1991; Bradford et al. 1993; Blaustein et al. 1994; Sjogren 1994;

Vos and Chardon 1998) and have revealed that a number of compounding effects such as inbreeding, environmental stochasticity and catastrophes, can quickly lead to local extinction.

Habitat loss and fragmentation and ensuing salinisation are recognized as primary concerns for species in Australia’s temperate regions and it is predicted that human impacts in the wheat-sheep zone of Australia will result in the extinction of 53% of all species (Morton 1999). The agricultural (wheatbelt) region of Western Australia

(WA) has experienced widespread habitat alteration for agriculture, involving the clearance of 140 000 km2 in less than 150 years (Saunders 1989). The bulk of native vegetation (54%) was cleared between 1945 and 1968 (Saunders et al. 1993). In most areas of the wheatbelt, less than 10% of the native vegetation now remains as small, isolated remnants (Hobbs 1993). Such rapid and widespread landscape change has had a great impact on endemic fauna and flora. Habitat fragmentation and salinisation caused by clearing are the key mechanisms responsible for the loss of plant and animal species in this region and will have ongoing future impacts (Hobbs and

Hopkins 1990).

Main (1990) suggested that the burrowing frog Heleioporus albopunctatus is declining in the WA wheatbelt as a result of increases in salinity. Main’s argument was based on museum records of this species, which showed that H. albopunctatus had not been collected in recent times from a number of wheatbelt shires, particularly those with the longest histories of land clearance and the largest areas affected by salinity. This thesis investigates the population structure and population dynamics of

H. albopunctatus to determine the likely current and future population structure of this species in a highly fragmented and rapidly degrading landscape.

Study species

Heleioporus albopunctatus is distributed widely throughout the semi-arid region of

Western Australia from near Kalbarri in the north-west to Esperance in the south-east.

Much of this distribution now coincides with the wheat-sheep farming region of

Western Australia. Accordingly, much of the range of this species has been subject to habitat loss and salinisation (eg. Saunders et al. 1993).

H. albopunctatus has terrestrial-egg deposition (Lee 1967). Males excavate burrows up to 1 m long in sandy substrates surrounding ephemeral waterbodies and commence calling in autumn (March/April). Amplexus occurs in the burrow and females deposit a foamy egg clutch in a chamber at the base of the burrow.

Embryonic development to mid-stage tadpoles takes place within the eggs, but successful hatching is dependent on winter rains filling the waterbody and flooding burrows. Tadpoles then hatch and complete development in the pond (Lee 1967).

Breeding populations of H. albopunctatus in the agricultural region of WA now consist of either large populations focused around a variety of anthropogenic and natural breeding ponds, or small, outlying populations, often in paddocks with no obvious surface water.

Study Objectives

In this thesis, I investigated the survival, ecology and long-term persistence of H. albopunctatus in the context of a highly man-modified environment – the central wheatbelt region of Western Australia. An examination of gene flow amongst populations can highlight the connectivity of populations in a metapopulation context and highlight any barriers to dispersal. I compared the genetic structuring of populations in the continuously vegetated Darling Scarp with those of the wheatbelt.

Consequently I was able to investigate the potential impacts of habitat fragmentation on gene flow across the east-west range of H. albopunctatus.

I examined the embryonic mortality of populations from across the central wheatbelt to gain essential autecological information and determine whether small populations in marginal habitats (those suffering from increased salinity and habitat fragmentation) were at risk from inbreeding (manifested as reduced reproductive fitness).

To gain an insight into the regional persistence of central wheatbelt populations of this species, I studied 48 breeding sites to gauge fecundity, larval survivorship and recruitment success over three years. My goal was to determine the importance of large versus small populations and gain an insight into the importance of source and possible sink populations in this landscape.

An intensive mark-recapture program allowed me to investigate dispersal, site fidelity and survivorship of adult frogs as well as estimate the size of populations.

Instead of an intensive 15 year single population study, I undertook a three year study at five sites. This gave a high level of replication though the duration was limited.

Finally, the total data package was integrated into a regional population viability analysis (PVA), using the life-history data acquired during this study. This allowed me to investigate the long-term impacts of identified threats on the persistence of this species over the next 100 years and to make management recommendations that may contribute to the persistence of this species at a landscape scale.

Chapter 2: The population genetic structure of the Western Spotted

Frog, Heleioporus albopunctatus (Anura: Myobatrachidae) in a fragmented landscape in south-western Australia.

Introduction

The predicted consequences of habitat fragmentation for many populations are a reduction in population connectivity, with subsequent reduction in gene flow towards isolated demes (Cunningham and Moritz 1998). Isolated populations with reduced dispersal capabilities may suffer erosion of heterozygosity and inbreeding depression, generating heightened genetic differences between populations (Lacy and

Lindenmayer 1995).

The consequences of habitat loss and modification such as genetic subdivision, inbreeding and local and regional extinction have been studied for wheatbelt mammals (Friend 1991), birds (Saunders 1989) and reptiles (Kitchener and How

1982; Sarre et al. 1995), but there have been no studies on wheatbelt amphibians.

Main (1990) suggested that the burrowing frog Heleioporus albopunctatus Lee (1967) is declining as a result of increases in salinity. Main’s argument was based on museum records of this species which showed that H. albopunctatus had not been collected in recent times from a number of wheatbelt shires, particularly those with the longest histories of land clearance and the largest areas affected by salinity.

This study consisted of an investigation into the broadscale patterns of genetic structuring in H. albopunctatus populations in the WA wheatbelt. The hypothesis being tested was that the impacts of habitat fragmentation on populations would be expressed as reduced gene flow across the wheatbelt range of H. albopunctatus in comparison to populations in continuous habitat. The importance of genetic

structuring in relation to habitat fragmentation, the maintenance of metapopulation structure and the conservation of local and regional populations are discussed.

Materials and Methods

Sampling methodology

Because of their burrowing habit adults are difficult to capture, so I collected tadpoles from 22 populations across the central wheatbelt region of WA (Fig. 2.1, Table 2.1).

The central wheatbelt is the oldest and most highly cleared agricultural region in the range of H. albopunctatus, and to determine broad-scale levels of genetic structuring, an approximate geographical transect was sampled from the continuously forested

Darling Scarp on the western edge of the species’ range, to the eastern edge of their range near Southern Cross. Twenty populations were spread across the central wheatbelt and two populations were sampled on the western edge of the range of H. albopunctatus (100 km east of Perth) in mixed marri (Corymbia calophylla) and wandoo (Eucalyptus wandoo) forest. The forest is not a replicate of the former wheatbelt environment and it may have been affected by logging and burning regimes but it represents the only remaining continuously-vegetated portion of the current range of H. albopunctatus.

Some populations of H. albopunctatus were located by listening for calling males during the breeding season (April-June) but most were found by netting for tadpoles at likely locations. Tadpoles were identified from descriptions in Lee (1967) and information supplied by M. Cowan (pers. comm.). Populations sampled included sites in large nature reserves and small isolated populations in paddocks and drains.

Populations were sampled with the aim of taking a minimum of 50 tadpoles for

analysis. To prevent complete loss of recruitment at small populations, ponds were netted to exhaustion and no more than half the standing crop of tadpoles were taken.

At sites where 50 tadpoles were easily collected, the complete number of tadpoles present in the pond was not counted. Sample sizes varied between 10 and 84 tadpoles

(mean = 40). Captured tadpoles were placed immediately into liquid nitrogen and stored at –80oC for 6 months before electrophoresis.

Tadpoles were mostly collected from sites known to contain large populations of

H. albopunctatus and can therefore be presumed to represent a random selection from a heterogenous tadpole population. However, it is possible that some small sites where few tadpoles were present, could represent a single or a small number of egg clutches. This possibility is considered in the discussion of the results.

Laboratory methodology

Allozyme electrophoresis was performed on homogenates of liver on cellulose acetate gels as described by Richardson et al. (1986). Twenty-two enzyme loci were initially screened, with thirteen of these exhibiting activities which could be scored reliably.

Of these thirteen presumptive loci, only four were found to be polymorphic (the most common allele occurring with a frequency of less than 0.95 in at least one population). The analysis is therefore restricted to these four loci as they are the only informative loci available. The enzymes stained, and E.C. numbers (cf. Murphy et al.

1996) were Fumarate hydratase (FUMHH, E.C.4.2.1.2), Glucose-phosphate isomerase (GPI, E.C. 5.3.1.9), Isocitrate dehydrogenase (IDH, E.C. 1.1.1.42) and

Phosphoglucomutase (PGM, E.C. 2.7.5.1).

Analyses

Measures of population subdivision were determined using Wright’s Fst value (Wright

1978). Both genetic distance and Fst can be used to investigate species subdivision.

However, in situations where mutation and long term evolutionary change are not important considerations (such as examinations of the genetic structuring across a species range), Fst is a more appropriate and sensitive measure than genetic distance, particularly at low levels of genetic divergence (Slatkin 1993; Roussett 1997).

Determinations of Fst and allele frequencies were assessed using the web version

(http://wbiomed.curtin.edu.au/genepop/) of GENEPOP (Raymond and Rousset 1995).

Fst values in GENEPOP were calculated for each locus and averaged over all four loci, using the methods in Weir and Cockerham (1984). Fst values were calculated over all populations and between all population pairs in the study area. To test for the significance of Fst values, a bootstrap analysis using 1000 simulations was conducted in GDA web-based software (http://lewis.eeb.uconn.edu/lewishare/software.html) for the analysis of discrete genetic data (Lewis and Zaykin 2001). Calculation of inbreeding coefficients (FIS) and tests for agreement with Hardy-Weinberg equilibrium were also calculated using GDA.

A multidimensional scaled plot of the Fst matrix for all population pairs was compiled in Statistica 5.5 for Windows using the multidimensional scaling module.

This provided a visual representation of the genetic distance between populations, based on a pairwise matrix of Wright’s Fst values using combined values across the 4 polymorphic loci. Multidimensional scaling permits a graphical representation in two- dimensional space of the relative genetic similarities of populations (Berry 2001).

Multidimensional scaling is preferable to UPGMA clustering as it can elucidate both hierarchical and non-hierarchical relationships among populations (Lessa 1990). This

is particularly relevant in studies such as this one where populations may contain genes from multiple origins (Lessa 1990).

A regression analysis was used to investigate isolation by distance from a matrix of pairwise Fst values and geographical distance.

Results

Genetic structuring

Allele frequencies for the four variable loci are shown in Table 2.2. GPI was the most polymorphic allele and PGM the least. Most individuals sampled were heterozygous for GPI and PGM and the majority of populations were fixed for the IDH (b) and

FUMH (b) alleles, or had these as the most common alleles. Percentage heterozygosity (H) averaged 13.4% and ranged from 3.3% to 34% for all populations

(Table 2.2). Inbreeding coefficients (f) ranged from 0 for H2 to 0.47 for CKL (Table

2.2). Three populations; HMD, TNR and YKR deviated significantly from Hardy-

Weinberg (H-W) equilibrium at the IDH locus and CKL deviated at the FUMH locus.

In each case, deviation from H-W equilibrium was due to a deficit of heterozygotes.

Three of the populations that deviated from H-W equilibrium (CKL, HMD, YKR) also had the highest inbreeding coefficients of all populations sampled (Table 2.2).

The average extent of genetic subdivision (averaged across all loci) between all populations at the broadest geographical scale was low (Fst = 0.087 SE ± 0.049

(p<0.05).

Pairwise estimates of Fst showed significant differences between some population pairs (Tables 2.3a and 2.3b). The central wheatbelt population RS showed the strongest genetic differentiation from other populations with an Fst range of between

0.144 (± 0.182) and 0.565 (± 0.481) when compared with all other populations. The variance for these values was high and the high Fst values may therefore reflect the low sample size for this population (n=12) the smallest sample available. Six other populations (MS, CKL, Q1, Q2, HMD and KR; Table 2.3) had high Fst values in comparisons with other populations. In total, 37.7% of pairwise comparisons of populations had Fst values greater than 0.1, with 16% of population pairs having an Fst value between 0.2 and 0.56. Fourteen of the 22 populations had one or more pairwise

Fst values that differed significantly from zero (Table 2.3). Seven populations, H1,

HW, KR2, MR, Q2, RS and DW all showed more than 3 significant differences with other populations. RS was the most differentiated population.

Multidimensional Scaling

The results of the multidimensional scaling plot are displayed in Fig. 2. The distribution of populations in two dimensions is suggestive of a random distribution, with three exceptions, populations HD, RS, YR2 and possibly MR (Fig. 2).

As part of the MDS plot, a stress value was also calculated. This value ranges between 0 (good) and 1 (poor), indicating how well the plot fits the original data

(Berry 2001). The stress value for this plot was 0.09, indicating a very good fit to the data.

Geographic Separation

A map showing the geographic distribution and drainage features surrounding all populations sampled is shown in Fig. 2.1. There was no separation of populations from the Perth Hills (H1 and H2) from those elsewhere. MC and KR are the two most

geographically isolated southern populations but did not appear substantially differentiated in the MDS plot (Fig. 2). The genetic basis for the separation of RS is based on higher frequencies of GPI (a), GPI (d) and FUMH (a) than any other population.

RS had a moderate level of inbreeding as shown by an f value of 0.18 (Table 2).

Divergence of the remaining outliers was not obvious from an examination of allele frequencies, although populations such as MR were fixed at three loci (FUMH B,

IDH B and PGM C).

The MDS plot does not generate groupings that correspond with the geographical locations of populations sampled: west (Perth Hills), central (Kellerberrin/Tammin) and east (Merredin and surrounds). An isolation by distance analysis (Manly 1985) showed there was no significant relationship between geographic distance and Fst, (r =

-0. 136, p = 0.34). These results suggest that factors other than isolation by distance are responsible for the observed patterns of genetic structuring.

Drainage Patterns

An examination of drainage patterns in relation to populations sampled (Fig. 2.1), clarifies some of the relationships observed on the MDS plot (Fig.2.2). Populations that were in relatively close geographical proximity to one another such as CKL and

Q1/Q2 were actually at the base of separate drainage systems. This may explain their genetic separation on the MDS plot (Fig. 2.2). Similarly, MS and YR2 (also close geographically but separate on the MDS plot) are on opposite sides of a drainage system.

Conversely, populations that were closely grouped on the MDS plot (such as Q2 and MS) were widely separated geographically and on completely unconnected drainage systems (Fig. 2.1).

The two populations in the Perth hills were geographically close, and on the same drainage system, yet they were genetically separated (Fig. 2.2). The outlying distribution of H1 is likely to be due to its greater heterozygosity at the GPI (b) and

(d) alleles than H2 that also lacks the GPI (a) and GPI (d) alleles. Thus drainage patterns clearly do not correlate with observed patterns of genetic structuring in many cases.

Discussion

The significant but low overall Fst for H. albopunctatus is consistent with moderate levels of either current or historical gene flow and dispersal throughout the central wheatbelt range of this species. The average Fst of 0.087 (± 0.049) is similar to that found in several other studies of genetic structuring in amphibian populations. Berry

(2001) reported a significant overall Fst of 0.088 in H. psammophilus and Seppa and

Laurilla (1998) reported Fst values of 0.068 and 0.019 respectively, in studies on Rana temporaria and Bufo bufo. Both studies concluded that high levels of gene flow existed between populations, though Fst is not necessarily a good surrogate for measures of gene flow (Whitlock and McCauley 1999). By contrast, studies on

Geocrinia species in southwest WA, demonstrated high genetic subdivision in

Geocrinia alba (Fst = 0.27) and G. vitellina (Fst = 0.302) with both species exhibiting extreme breeding site philopatry (Driscoll et al. 1994; Driscoll 1998). Geocrinia species studied by Driscoll (see above) have specialised breeding biologies and restricted habitat availability, favouring the evolution of philopatry (Wardell-Johnson

and Roberts 1993; Roberts et al. 1999). However, in Rana temporaria (Fst = 0.388) and Bufo bufo (Fst = 0.487) in England, Hitchings and Beebee (1997; 1998) suggested that subdivision in these species was generated by habitat fragmentation in urban environments. In the present study, the lack of differentiation between populations in continuous habitat (H1 nad H2) and wheatbelt populations as well as the weak relationship between Fst and geographic distance, indicate that habitat fragmentation is not a leading cause of genetic structuring in H. albopunctatus. It should be recognised that allozymes may not detect all genetic variation and more rapidly evolving markers, e.g., microsatellites, may give a more sensitive view of current patterns of dispersal. If data in this study represent historical patterns of dispersal then they represent a minimum model of dispersal to be maintained in current population management.

The percentage heterozygosity (%H) of H. albopunctatus populations was relatively high (3.3 – 34%) and the inbreeding coefficient (Fis) variable (0.028 – 0.47).

Other studies of amphibian populations have reported comparatively low levels of heterozygosity: e.g. 0.04 – 0.21% for Pseudophryne corroboree (Osborne and

Norman 1991); 1.7 – 3.5% for Rana temporaria and 3.3-10.6% for Bufo bufo

(Hitchings and Beebee 1997; 1998); 0-4% for R. temporaria (Reh and Seitz 1990) and

0-9.3% for Geocrinia alba and G. vitellina (Driscoll 1998). Data on Fis are scarce but

Hitchings and Bebee (1997) reported high inbreeding coefficients for Rana temporaria (Fis = 0.539 – 0.582). Reported values of %H and Fis for H. albopunctatus, in tandem with a low overall Fst, may be consistent with dispersal occurring between populations, with sufficient frequency to maintain the heterozygosity of most populations (Table 2.2).

Although the dispersal capacity of H. albopunctatus is unknown, the congener,

H. eyrei is known to move up to 2 km from breeding sites (Bamford 1992). Research

(Chapter 5) on dispersal and population demographics of H. albopunctatus, suggests some, but limited adult dispersal but no data are available on dispersal by tadpoles or metamorphs. Populations sampled in the Perth hills were not more or less genetically differentiated than populations in the central wheatbelt (Fig. 2.2) and also had similar levels of heterozygosity and inbreeding coefficients to central wheatbelt populations, suggesting the possibility of widespread gene flow between central wheatbelt and hills populations. These genetic data suggest that H. albopunctatus may have the dispersal capacity required to persist in a fragmented landscape (but see comments above on current versus historical patterns of gene flow).

Despite low overall levels of genetic differentiation, there are still three populations that are significantly differentiated from others (RS, HD, YR2; Fig. 2.2).

All three have a large number of significant, high pair-wise Fst values and were substantially genetically differentiated from all other populations (Table 2.2; Fig. 2.2).

The processes generating these differences are difficult to establish as.these populations were not geographically more isolated than other central wheatbelt populations (Fig. 2.1). These significant differences in genetic structuring between some populations, suggest either small population effects, e.g. genetic drift, inbreeding or small sample sizes, or, the presence of subtle barriers to gene flow.

Physical barriers to gene flow generated by habitat fragmentation may also be contributing to observed patterns of genetic structuring by blocking normal dispersal routes: e.g. salt affected drainage systems may reduce tadpole dispersal, or large areas of cleared terrain may inhibit dispersal into isolated populations.

There is the possibility that tadpoles sampled from some sites may have arisen from a single egg clutch. Whilst this possibility cannot be excluded, it remains unlikely as sites sampled were chosen as representative of good natural breeding sites for this species. Populations with a similar number of breeding burrows sampled during an intensive population study in Kellerberrin, WA (Chapter 3) were estimated to average at least 25 egg masses. Furthermore, the mortality rate from egg to tadpoles averaged 98% and thus a small number of tadpoles present in a pond may still be representative of a large number of clutches (Chapter 3).

The MDS plot (Fig. 2.2) and drainage maps (Fig. 2.1) revealed that some of the outlying populations were geographically separate from other populations (e.g. MC,

RS), but that was not necessarily the case (eg. KR). There was no pattern of genetic differentiation simply related to drainage systems including hyper-saline systems such as the Salt River, where populations on opposite side of the drainage (eg. MR vs MS) were genetically similar to one another (Fig. 2.1). Some of the most differentiated populations (e.g. RS and KR, Table 2.2) also had the smallest sample sizes.

The general picture therefore seems to be one of populations connected by moderate levels of gene flow but showing significant (although low) levels of subdivision overall and characterised by high levels of differentiation between some population pairs and inbreeding in some sites.

The fact that only four loci in this study exhibited scoreable activity may have affected the resolution of some analyses. Gorman and Renzi Jr. (1979) discussed the reduced resolution in allozyme estimates of genetic distance and heterozygosity with few polymorphic loci. In this study, these impacts were unavoidable due to the absence of further polymorphic loci (22 were screened). Future studies utilising modern DNA techniques are recommended and may be useful for comparison.

Metapopulation models for H. albopunctatus.

Metapopulation models outlined by Harrison (1991) are associated with high levels of dispersal and therefore gene flow (classic and patchy models) or, low levels of dispersal and high differentiation (non-equilibrium models). The source-sink model of Harrison (1991) is characterised by a mixture of successfully recruiting sites and a number of temporally fluctuating satellite sites characterised by high rates of extinction and unidirectional dispersal from source to sink. It might be expected, therefore, that there would be a pattern of high genetic subdivision amongst unrelated source populations with varying levels of relatedness amongst source-sink populations, unrelated to geography. This is consistent with the overall pattern of genetic structuring observed but I cannot eliminate other models given that I do not have data on relatedness of tadpoles in the samples in some differentiated populations.

Direct data on recruitment and dispersal patterns for all life history stages are needed to test this prediction about metapopulation structure.

Recognising the mechanism for dispersal, adult, metamorphs or tadpoles, and the impact the cleared matrix may have on that dispersal are critical steps in ensuring regional persistence of frogs such as H. albopunctatus. Distinguishing current from historical patterns of dispersal will also be critical to ongoing management of these populations.

Table 2.1. Locality and size of breeding populations sampled for genetics.

Population UTM UPS Location Sample Size

(WGS 84)

MM 50623089E Mount Moore 70

6547065N

BR 50595074E Billyacatting Rock 64

6565046N

CH 50637029E Chingah Hills 74

6495041N

HD 50623257E Hunt’s Dam 36

6520351N

TNR 50634072E Tandegin Rock 84

6502074N

TR 5061546E Totadgin Rock 76

6520351N

MS 50556092E Mount Stirling 50

6477758N

YR2 5056916E Yoting Rock 2 54

6469981N

DW 50568684E Deep Well Rd 24

6512292N

RS 50562973E Roger Scott’s property 12

6499860N

Table 2.1. (cont.).

CKL 50556051E Creekline in 30

6477728N Woollering N.R.

Q1 50569551E Quarry site 1 16

6513741N

Q2 50569551E Quarry site 2 50

6513741N

MR 50560114E McNeil Rd 22

6488686N

HMD 50571377E Population study site 42

6522906N three

MC 50662014E Mount Cramphorne 28

6477315N

KR2 50567399E Kokerbin Rock 2 28

6472528N

KR 50566241E Kokerbin Rock 10

6472176N

HW 50546772E Hunt’s Well 22

6496202N

YKR 50548383E Yorkrakine Rock 56

6523246N

H2 50456057E Hills site 2 14

6428063N

H1 50456049E Hills site 1 50

6429910N

Table 2.2. Allele frequencies for 22 populations of H. albopunctatus for the 4 variable loci detected. Sample sizes (N), inbreeding coefficients (f) and mean percentage heterozygosity (%H) with standard error (SE) are also shown.

Popn MM BR CH HD TNR TR MS YR2 DW RS CKL Q1 Q2 N 70 64 74 36 84 76 50 54 24 12 30 16 50

GPI A 0.000 0.000 0.014 0.000 0.012 0.071 0.020 0.019 0.042 0.250 0.000 0.125 0.060 B 0.014 0.062 0.081 0.056 0.048 0.024 0.100 0.019 0.167 0.000 0.000 0.000 0.000 C 0.886 0.766 0.784 0.889 0.821 0.810 0640 0.889 0.667 0.333 0.767 0.688 0.680 D 0.100 0.141 0.122 0.056 0.107 0.095 0.240 0.074 0.125 0.417 0.233 0.062 0.260 E 0.000 0.031 0.000 0.000 0.012 0.000 0.000 0.000 0.000 0.000 0.000 0.125 0.000

PGM B 0.000 0.103 0.027 0.000 0.131 0.026 0.200 0.024 0.042 0.111 0.000 0.125 0.200 C 1.000 0.897 0.973 1.000 0.869 0.974 0.800 0.976 0.958 0.889 1.000 0.875 0.800

IDH A 0.000 0.000 0.000 0.000 0.048 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 B 1.000 1.000 1.000 1.000 0.952 1.000 1.000 1.000 0.909 1.000 1.000 1.000 1.000 C 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.091 0.000 0.000 0.000 0.000

FUM A 0.000 0.000 0.048 0.000 0.012 0.000 0.000 0.000 0.000 0.500 0.000 0.000 0.000 H B 1.000 1.000 0.952 1.000 0.988 1.000 1.000 1.000 1.000 0.500 0.600 1.000 1.000 C 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.400 0.000 0.000 f 0.099 0.16 0.079 0.062 0.028 0.11 0.22 0.016 0.23 0.18 0.47 0.17 0.30 %H 6.8 14 8.1 5.1 15 9.5 21 6.3 19 34 21 18 20 SE 0.052 0.093 0.085 0.031 0.065 0.088 0.13 0.048 0.11 0.14 0.12 0.12 0.12

Table 2.2 (cont.).

Popn MR HMD MC KR2 KR HW YKR H2 H1

N 22 42 28 28 10 22 56 14 50

Pgi A 0.000 0.000 0.000 0.000 0.000 0.000 0.036 0.000 0.020 B 0.000 0.024 0.000 0.000 0.000 0.000 0.018 0.071 0.180 C 0.773 0.952 0.643 0.964 0.400 0.818 0.946 0.929 0.700 D 0.227 0.000 0.321 0.000 0.500 0.182 0.000 0.000 0.100 E 0.000 0.024 0.036 0.036 0.100 0.000 0.000 0.000 0.000

Pgm B 0.000 0.095 0.000 0.036 0.000 0.091 0.015 0.000 0.000 C 1.000 0.905 1.000 0.964 1.000 0.909 0.985 1.000 1.000

Idh A 0.000 0.143 0.107 0.000 0.000 0.000 0.056 0.000 0.000 B 1.000 0.857 0.893 1.000 1.000 1.000 0.944 1.000 1.000 C 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000

Fum A 0.000 0.000 0.143 0.000 0.000 0.000 0.028 0.000 0.000 B 1.000 1.000 0.857 1.000 1.000 1.000 0.972 1.000 1.000 C 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 f 0.25 0.46 0.29 0.052 0.017 0.13 0.36 0.00 0.18 %H 8.8 13 23 3.5 14 12 7 3.3 12 SE 0.088 0.053 0.10 0.017 0.145 0.072 0.019 0.033 0.12

Table 2.3a. Pairwise Fst estimates for all population pairs. Significant Pairwise differences at the 5% nominal level, are shown in bold.

MM BR CH HD TNR TR MS YR2 DW RS CKL Q1 Q2 MR HMD MC KR2 KR HW YKR H2 H1 MM 0 BR 0.06 0 CH 0.12 0.01 0 HD 0.23 0.16 0.18 0 TNR 0.05 0.08 0.12 0.23 0 TR 0.07 0.05 0.05 0.11 0.04 0 MS 0.12 0.02 -0.01 0.18 0.12 0.05 0 YR2 0.33 0.16 0.11 0.19 0.39 0.18 0.09 0 DW 0.04 0.12 0.15 0.25 0.00 0.14 0.16 0.51 0 RS 0.43 0.25 0.27 0.2 0.51 0.31 0.25 0.19 0.54 0 CKL 0.07 0.01 0.04 0.13 0.07 0.02 0.02 0.14 0.14 0.26 0 Q1 0.06 0.00 0.02 0.19 0.03 0.01 0.02 0.17 0.06 0.36 0.00 0 Q2 0.08 0.12 0.13 0.29 -0.01 0.03 0.14 0.43 0.03 0.57 0.09 0.04 0 MR 0.14 0.07 0.07 0.11 0.14 0.02 0.07 0.05 0.19 0.14 0.04 0.07 0.16 0 HMD 0.04 0.09 0.12 0.22 -0.02 0.05 0.12 0.42 0.01 0.48 0.06 0.03 -0.01 0.13 0 MC 0.07 0.02 0.07 0.19 0.00 0.01 0.08 0.26 0.04 0.41 0.02 0.01 0.01 0.09 0.00 0 KR 0.02 0.00 0.03 0.19 0.02 0.01 0.03 0.20 0.04 0.35 0.01 0.00 0.04 0.08 0.02 0.01 0 KR2 0.07 0.03 0.06 0.18 0.01 -0.01 0.06 0.20 0.05 0.36 0.00 0.00 0.02 0.06 0.01 0.00 0.01 0 HW 0.00 0.11 0.16 0.24 0.01 0.07 0.17 0.44 0.00 0.50 0.09 0.07 0.03 0.16 0.00 0.03 0.04 0.04 0 YKR 0.04 0.01 0.01 0.16 0.01 -0.01 0.02 0.20 0.07 0.33 0.02 -0.02 0.02 0.06 0.03 0.00 -0.02 0.00 0.06 0 H2 0.00 0.08 0.12 0.19 -0.03 0.09 0.11 0.44 -0.02 0.42 0.06 0.03 0.00 0.13 -0.04 0.01 0.01 0.01 -0.04 0.04 0 H1 0.12 0.05 0.09 0.21 0.07 0.03 0.06 0.2 0.12 0.38 -0.02 0.02 0.08 0.09 0.05 0.03 0.04 0.01 0.11 0.04 0.05 0

Table 2.3b. Corresponding pairwise variances for Table 3a. MM BR CH HD TNR TR MS YR2 DW RS CKL Q1 Q2 MR HMD MC KR2 KR HW YKR H2 H1

MM 0 BR 0.04 0 CH 0.09 0.01 0 HD 0.21 0.16 0.18 0 TNR 0.02 0.03 0.06 0.16 0 TR 0.04 0.03 0.03 0.10 0.01 0 MS 0.10 0.01 -0.01 0.19 0.07 0.04 0 YR2 0.27 0.13 0.09 0.20 0.20 0.09 0.08 0 DW 0.01 0.05 0.09 0.18 0.00 0.04 0.10 0.28 0 RS 0.53 0.33 0.32 0.27 0.47 0.35 0.32 0.24 0.55 0 CKL 0.05 0.01 0.03 0.13 0.03 0.01 0.02 0.10 0.06 0.33 0 Q1 0.04 0.00 0.02 0.16 0.01 0.00 0.01 0.12 0.03 0.39 0.00 0 Q2 0.03 0.04 0.07 0.16 0.00 0.01 0.08 0.19 0.01 0.48 0.03 0.02 0 MR 0.11 0.06 0.06 0.11 0.08 0.01 0.06 0.04 0.12 0.18 0.04 0.06 0.08 0 HMD 0.02 0.04 0.08 0.16 0.00 0.02 0.08 0.21 0.00 0.49 0.03 0.01 0.00 0.08 0 MC 0.03 0.01 0.04 0.15 0.00 0.00 0.05 0.16 0.01 0.41 0.01 0.00 0.00 0.06 0.00 0 KR 0.01 0.00 0.02 0.17 0.01 0.01 0.02 0.16 0.02 0.40 0.01 0.00 0.02 0.06 0.01 0.01 0 KR2 0.04 0.02 0.04 0.14 0.00 0.00 0.04 0.13 0.02 0.36 0.00 0.00 0.01 0.04 0.00 0.00 0.01 0 HW 0.00 0.05 0.10 0.17 0.00 0.03 0.11 0.26 0.00 0.50 0.04 0.03 0.01 0.10 0.00 0.01 0.02 0.02 0 YKR 0.02 0.01 0.01 0.13 0.00 0.00 0.02 0.13 0.02 0.39 0.01 -0.01 0.01 0.04 0.01 0.00 0.01 0.00 0.02 0 H2 0.00 0.04 0.09 0.15 -0.01 0.03 0.09 0.26 0.00 0.51 0.03 0.02 0.00 0.10 -0.01 0.00 0.01 0.00 0.01 0.02 0 H1 0.07 0.03 0.06 0.17 0.03 0.02 0.04 0.12 0.05 0.4 -0.01 0.01 0.03 0.06 0.02 0.01 0.03 0.00 0.05 0.02 0.02 0

Figure 2.1. Location of H. albopunctatus genetic sampling sites in south-west WA (see inset) and the central wheatbelt region (main map). Hatched lines indicate the extent of major saline drainages in the region.

2.5

1.5

Q2 0.5 H1 TNR MC RS HMD KR2 TR DW CKL MR 0 H2 HW YKR Q1 KR BR -0.5 MS MM CH YR2 -1

-1.5 -1.5 -1 -0.5 0 0.5 1 1.5 2 2.5 3

Figure 2.2. Multidimensional scaled plot of pairwise Fst differences between populations of H. albopunctatus.

Chapter 3: Embryonic Survival and Egg Numbers in Small and Large Populations of the Frog Heleioporus albopunctatus in Western Australia

Introduction

Frankham (1997) showed that extinction risks were much higher in small populations of many animal and plant groups. There are several possible explanations for this result including inbreeding depression, stochastic loss of small populations or because small populations occur in marginal habitats.

The impacts of inbreeding on fitness traits is uncertain (Lande 1988; Shields 1993).

Naturally small populations and large populations recently reduced in size may suffer from inbreeding depression and consequently reduced fitness (Hedrick and Miller

1992). Reed and Frankham (2003) reviewed 34 studies that all reported a loss of heterozygosity associated with reduced fitness. There are also contrary results, however

(e.g., Gundersen et al. 2001).

A common impact of inbreeding is on traits associated with egg or juvenile fitness.

For example, Daniels and Walters (2000) found that inbreeding reduced hatching success, fledgling survival and recruitment of red-cockaded woodpeckers. Keller (1998) found lower hatching rates of eggs from inbred female song sparrows. There have been few studies of inbreeding and its impact on embryonic survival in anurans. Hitchings and Beebee (1997; 1998) examined urban and rural anuran populations and found evidence of reduced fitness in inbred urban populations of Rana temporaria and Bufo bufo. Edenhamn et al. (2000) studied the impacts of inbreeding on hatching success and pre-metamorphic survival of Hyla arborea, and found that early (post-hatching) larval survival was lower in isolated breeding ponds but that egg survival to hatching was not lower than in other parts of the species’ range.

In addition to inbreeding effects, developing amphibian embryos are exposed to various natural hazards such as predation (Howard 1978; Miaud 1993), desiccation

(Humphries 1979) and infection by pathogenic fungi (Williamson and Bull 1994).

Abiotic factors including water depth and temperature (Seale 1982) and oxygen availability and uptake may also be critical determinants of embryonic survival (Licht

1970; Seymour and Roberts 1995). Breeding populations of amphibians may be small because some or any of these factors are more extreme than in other populations.

Marginal habitats may also limit energy available to breeding females, reducing clutch size.

Thus in small populations there are two confounded factors, habitat suitability and inbreeding, that may depress fitness relative to larger populations through increased embryonic mortality and possibly smaller clutch sizes. In this chapter I investigated the relationship between population size and fitness of H. albopunctatus populations in the semi-arid, agricultural regions of south-western Australia known as the wheatbelt (cf.

Saunders et al. 1993). I expect some small populations may be inbred because Chapter 2 established that some small populations of this species exhibit higher levels of inbreeding than larger populations, measured by heterozygosity and inbreeding coefficients ranging from He = 3.3%, f = 0.00 (n = 14) to He = 13%, f = 0.46 (n = 42) for 22 populations. Small populations may represent marginal habitats because successful breeding is rare in sites with low numbers of calling males and is associated with short hydro-periods and elevated salinity (Chapter 4).

Fitness was measured by scoring clutch size and embryonic survival to test the hypothesis that smaller populations had both lower embryonic survival and smaller clutch sizes. I also investigated the relationship between female body size and clutch

size, to control for any female body size effects, and developmental stage, to control for age-related embryonic mortality.

Methods

Egg Survival and Population Size

To estimate population size and investigate its relationship with embryonic mortality, I examined 16 H. albopunctatus breeding sites in the wheatbelt region of Western

Australia in 2000 and 2001. Breeding sites sampled covered most of the east-west range of this species. Suitable locations were examined for breeding burrows and all burrows at a site were excavated to obtain egg clutches. Excavations were sometimes thwarted by obstructions in burrows such as subterranean rocks or hard clay, as noted in Lee

(1967). Consequently, some egg clutches may have been missed but missed egg masses were presumed to represent a random sample of available egg masses. Only 37 % of excavated burrows contained egg masses (Chapter 4) which is also characteristic of the congener H. eyrei (Roberts unpublished data).

Excavated clutches were preserved immediately in the field in 10% buffered formol saline. In the laboratory, clutches were examined under a dissecting microscope with strong white light. Each egg clutch was examined for unfertilized eggs (small and undeveloped, presumably not fertilized), or dead eggs (normal sized but milky and opaque with only the capsule remaining). Dead and infertile eggs were combined as dead eggs in analyses reported below.

Burrows were counted at each site sampled, and used as an index of local population

2 size. H. albopunctatus burrow counts are related significantly (r = 0.507, F1,13 = 12.35, p = 0.004, n = 15) to actual population size (minimum number known to be alive from a

comprehensive trapping programs at five sites over three years – Chapter 5) by the following equation: population size = 0.766 x number of burrows. To investigate variation in egg survival in relation to population size, the average number of dead/ infertile eggs at each site was regressed against the number of burrows at that site.

Egg Survival and Developmental Stage

I staged each excavated egg clutch according to the modified Gosner (1960) stages of

Anstis (2002). I plotted percentage embryonic mortality against stage to detect discontinuities in development and survival. Gosner stage is an ordered, categorical scale and is not suitable for simple regression analysis. I also compared mortality in clutches designated as early (Gosner stages 12-20) or late (Gosner stages 21-28) in development using a t-test.

Clutch Size and Population Size

To investigate whether smaller populations had a reduced clutch size, I regressed the number of eggs/clutch against population size (as determined by the number of burrows). I counted the total number of eggs per clutch in the egg masses excavated from each site.

Female Body Size and Clutch Size

To investigate the relationship between female body size and clutch size, I dissected out and counted the eggs from all (6) gravid female H. albopunctatus available in the West

Australian Museum collection and two other specimens available from accidental field mortalities. I averaged four snout-vent length (SVL) measurements taken with digital

calipers and used regression analysis to explore the relationship between mean female

SVL and clutch size.

Results

I collected a total of 55 egg clutches from 16 sites (each with at least one egg clutch) across the central wheatbelt of WA. “Population size” ranged from two to thirty burrows and the number of excavated egg clutches per site ranged from one to eight.

Egg Survival and Population Size

Clutch size ranged from 42 – 876 eggs. Mean clutch size (n = 55) was 391.2 ± 21.17

(standard error). Across all sites, 80% of egg clutches had some egg mortality but only a small percentage of the eggs in each clutch were dead, with average egg mortality 11.25

(± 3.03) eggs/clutch (2.80 ± 0.65 expressed as a % range 0-21.17%). A regression of egg mortality against the number of breeding burrows for all populations was not significant (Fig.3.1). Four of the 55 egg clutches (7.27%) displayed signs of infestation by larvae of the phorid fly Aphiura breviceps, indicated by the presence of larvae and adult female flies throughout the egg clutch.

Egg Survival and Developmental Stage

Egg clutches ranged from stage 12 to stage 30, but most were at stages 24-28. There was no obvious relationship between percentage egg mortality and developmental stage but the individual clutches with the highest mortality were at the latest stages of development sampled (Fig. 3.2).

Clutch Size and Population Size

Clutch size was regressed against burrow counts, but no significant relationship was

2 detected (r = 0.00, F1,53 = 0.018, p = 0.89).

Female Body Size and Clutch Size

The 8 specimens available encompassed almost the full range of body sizes (46.6 - 79.2 mm) reported for breeding females by Tyler et al. (2000). A regression of clutch size on

2 SVL found no significant relationship (r = 0.02, F1,6 = 0.13, p = 0.73). I assumed that museum specimens had a complete clutch because all eggs were still in the ovary and had not been released into the oviduct.

Discussion

Clutch Size and Population Size

I found no evidence of a smaller clutch size in small populations. In this study, the number of eggs per clutch ranged from 42 – 876. Lee (1967) reported a range of 400-

600 eggs/clutch for Heleioporus albopunctatus (N = 9) and Tyler (1994) reported a range of 250 – 700 eggs/clutch (based on Lee 1967 and Davies 1991). Sample sizes were small for the earlier estimates and the greater range reported here is likely to be a reflection of the greater sample size in this study. The lower minimum clutch size reported here may be the result of split clutches. Lee (1967) stated that H. albopunctatus females are likely to deposit only one clutch per breeding season and observed no evidence of split clutches. The breeding biology of H. albopunctatus is difficult to study

in-situ (and ex-situ cf. Lee 1967), but the presence of multiple clutching or clutch splitting as found in Afrixalus delicatus (Backwell and Passmore 1990) and

Pseudophryne spp Woodruff (1976a) could only be shown by following individual females or by genetic analysis.

Female Body Size and Clutch Size

No correlation was found between female SVL and clutch size. Differences in clutch size related to body size can therefore be eliminated as a confounding variable when considering the relationship between clutch size and population fitness, though I accept that the sample size justifying this conclusion is small. However, these data are consistent with patterns in other Australian frog species where only 3 of 11 species showed a positive relationship between clutch and body size (Dziminski 2000).

Egg Survival and Population Size

I detected no evidence of reduced egg survival in small as opposed to large populations.

This contradicts some other studies on frogs. Hitchings and Beebee (1997; 1998) found reduced survivorship and increased levels of developmental abnormality in small, less heterozygous populations of Bufo bufo and in small, genetically subdivided populations of Rana temporaria. Edenhamn et al. (2000) found no differences in egg survival of

Swedish Hyla arborea populations that displayed lower levels of genetic variation than those in the rest of Europe. I found small isolated populations of H. albopunctatus, some with moderate to high levels of apparent inbreeding, detected from genetic data

(Chapter 2). There was, however, overall gene-flow and a limited pattern of subdivision consistent with frequent dispersal. This has also been observed in other Heleioporus

species (Berry 2001). The low rate of embryonic mortality coupled with the presence of overall gene flow between populations, may indicate that inbreeding does not have a significant impact on this species. Egg mortality in H. albopunctatus was low, with an average of only 2.8% egg mortality in each clutch at the time of examination. This compares favorably with many other species reported in the literature (Table 3.1).

Few data on egg mortality are available for species that deposit eggs in terrestrial burrows but have an aquatic tadpole stage. Woodruff (1976b) reported embryonic mortality of the ground nesting Pseudophryne bibroni, P. dendyi and P. semimarmorata at less than 5% on average (range 0-11%). These data are comparable with H. albopunctatus in this study (2.8% average egg mortality), although the range was greater for H. albopunctatus (0-21.2%) than for Pseudophryne spp.

Egg Survival and Developmental Stage

No relationship was found between developmental stage and egg mortality. Packer

(1966) reported that eggs of the congeneric H.eyrei never hatched earlier than stage 21 –

22, but could hatch much later, up to stage 29. It is possible that mortality events that occurred during early stages may not be detected at a late stage (e.g. if dead eggs were digested completely by fungi), but since complete egg capsules containing decayed embryos were clearly detected, this is unlikely to be an issue. Woodruff (1976b) observed that most egg mortality in Pseudophryne species occurred during gastrulation

(Gosner stage 10-12). Further examination of early stage egg clutches would be required to confirm the relationship between Gosner stage and mortality, but in this study I rejected the hypothesis that embryonic mortality was related to developmental stage.

Causes of Egg Mortality

Predation is a well-documented source of egg mortality in anuran pre-metamorphic survival studies. Rogowitz et al. (2001), determined that predation accounted for 56% of the observed mortality in clutches of Eleutherodactylus cooki, but Williamson and

Bull (1994) found egg predation to be unimportant in Crinia signifera. Predation by leeches (Humphries 1979) and carabid beetles (Ehmann and Swan 1985) has been documented. Foxes excavated several H. albopunctatus burrows and this also affected egg survival, although the impact of vertebrate predators is poorly known. The incidence of predation on H. albopunctatus is likely to be greatly reduced by the relative inaccessibility of eggs in burrows.

Predation by larvae of a number of dipteran fly families can have a large impact on anuran egg survival. Lips (2001) documented predation of Hyla calypsa eggs by a drosophilid fly and dipteran predation accounted for 17.5 – 61% of egg death in four species of Hyperolius (Vonesh 2000). Phorid fly predation also inflicted significant egg mortality in Agalychnis annae (Villa and Townsend 1983). Four of 55 H. albopunctatus egg masses in this study were infested with flies and larvae of the phorid fly Aphiura breviceps, but further experimental work is required to determine the exact contribution of dipteran predation to egg mortality in this species (Davis and Disney 2003).

Fungal infection was a leading cause of egg mortality in studies of various anurans by Woodruff (1976b), Malone (1985), Williamson and Bull (1994) and Warkentin et al.

(2001); however, no obvious hyphae were observed in dead H. albopunctatus eggs and it was not possible to demonstrate the presence of fungi in egg clutches of this species.

None of the factors discussed above seem to be affecting small populations of H. albopunctatus any more than large populations. In summary, embryonic survival in H. albopunctatus was high, possibly reflecting the protective effect of depositing eggs in a

deep burrow, though the impact of phorid flies is still unknown (Davis and Disney

2003). I rejected the hypothesis of enhanced mortality or reduced clutch size in small populations, suggesting that inbreeding and marginal habitats are not affecting these characters.

Table 3.1: Reported embryonic mortality in the anuran literature classified according to egg deposition mode. * indicates that egg mortality estimates are based on laboratory studies. (Modified from Conroy, 2001).

Species Egg Deposition % Egg Reference

Mode mortality

Alytes obstetricans Parental Care 8.6-10.1 Reading and Clarke, 1988

Centrolenella colymbiphyllum Arboreal 3.3-44 McDiarmid, 1978

C. fleischmanni Arboreal 29-39 Greer and Wells, 1980

C. valerioi Arboreal 3.4-14.3 McDiarmid, 1978

Hyla calypsa Arboreal 73.3-78 Lips, 2001

Phyllomedusa guttata Arboreal ~6 Lutz, 1947

Polypredates leucomystax Arboreal 34 Yorke,1983

Bufo calamita Aquatic 14-54 Banks and Beebee, 1988

Crinia signifera Aquatic 32.4-98.3 Williamson and Bull, 1994

Hyla rosenbergi Aquatic 0-100 Kluge, 1981

Limnodynastes tasmaniensis Aquatic 90 Humphries, 1979

Limnodynastes tasmaniensis Aquatic 2.4-99 Roberts, 1993

Litoria ewingi* Aquatic 3-6 Watson and Martin, 1968

Litoria verreauxi* Aquatic 21-26 Watson and Martin, 1968

Pseudacris triseriata Aquatic 12.6-62.3 Kramer, 1978

Rana aurora Aquatic 8-9 Licht, 1974

Table 3.1 (cont.).

Rana macrocnemis Aquatic 12.6-38.8 Tarkhnishvili and

Gokhelashvili, 1999

R. pretiosa Aquatic 26-100 Licht, 1974

R. sylvatica Aquatic 3.4-7.5 Seigel, 1983

R. sylvatica Aquatic 13.7-58.4 Waldman, 1982

Geocrinia victoriana/laevis Terrestrial 0- 80.6 Gollman and Gollman, 1994

Heleioporus albopunctatus Terrestrial 0-21.2 This study

Philoria frostii Terrestrial 74 Malone, 1985

Pseudophryne bibronii* Terrestrial 2.8-5.7 Woodruff, 1976b

P. dendyi* Terrestrial 4.1-8.3 Woodruff, 1976b

P. semimarmorata* Terrestrial 1.1-6.7 Woodruff, 1976b

Chirixalus eiffingeri Tree Hollow 6-86 Kam et al., 1997

Figure 3.1. The relationship between the number of dead eggs per clutch and

2 population size (r = 0.0023, F1,53 = 0.12, p = 0.73).

Figure 3.2. Percentage embryonic mortality in relation to Gosner stage.

Chapter 4: Larval recruitment success of the frog Heleioporus albopunctatus in a West Australian agricultural landscape.

Introduction

Research on amphibian declines has focused attention on the need for a thorough understanding of metapopulation dynamics, particularly those factors regulating survival and recruitment (Cecil and Just 1979; Alford and Richards 1999). Success at the population level is determined primarily by the number and quality of metamorphosing larvae successfully recruited into the adult population (Semlitsch

2000) but this may be modified by dispersal. In a landscape context, juvenile amphibians may be primarily responsible for dispersal and gene flow between geographically separated populations affecting both demographic and genetic properties of populations (cf. Gill 1978; Breden 1987). Ongoing recruitment is therefore essential for the maintenance of regional metapopulations.

Recruitment to the adult stage is regulated at several stages of the amphibian life cycle (Wilbur 1980), but mortality is usually greatest at the larval stage (Istock 1967;

Wilbur 1980). Studies on larval recruitment have shown success rates that are often less than 1% from egg to metamorphosis (Herreid and Kinney 1966; Licht 1974; Semslitch

1987) and these studies point to the importance of factors in the larval rather than the embryonic phase of the lifecycle as regulating adult recruitment.

The successful recruitment of larvae to metamorphosis is determined by several factors including hydroperiod (eg. Pechmann et al. 1989), pond size and depth

(DiMauro and Hunter 2002), competition (Newman 1988; Pfennig 1990; Wilbur 1987) and predation (Licht 1974; Cecil and Just 1979; Tyler et al. 1998). Hydroperiod and

pond characters may exert a particularly strong influence on larval survival and subsequent recruitment success. While many landscapes include large permanent lakes and rivers, temporary pools and ponds are often the most numerous type of wetland in natural landscapes (Gibbs 1993; Semlitsch and Brodie 1998). Temporary wetlands are used by more species and recruit more metamorphs than larger, more permanent, pools and ponds (Pechmann et al. 1989; Semlitsch et al. 1996; Semlitsch 2000). These ephemeral wetlands are subject to greater fluctuations in hydroperiod, depth, area and water-quality characteristics, and these may exert a greater influence on recruitment success than biotic factors such as predation and competition (cf. Caldwell 1987;

Berven 1995; Tinsley and Tocque 1995; Tarkhnishvili and Gokhelashvili 1999;

DiMauro and Hunter 2002). Abiotic and biotic pond-related factors may reduce breeding efforts and metamorph fitness in certain ponds, resulting in ponds acting as

“ecological traps” (DiMauro and Hunter 2002).

The wheatbelt region of Western Australia is situated in the semi-arid south-west of the state within the 250-500 mm rainfall zone. As a result of the semi-arid climate, there are few permanent natural wetlands available for amphibian breeding activity and most of these have become saline as a result of widespread vegetation clearance (Hobbs and

Hopkins 1990; Main 1990). Consequently, frogs have had to contend with a human- modified landscape comprising a diversity of ponds, differing greatly in their biotic and abiotic composition.

The objective of this chapter was to compare the recruitment success of populations of Heleioporus albopunctatus breeding in a diverse array of mostly anthropogenic, ephemeral pools. Specifically, I wanted to 1) document patterns of successful recruitment to the metamorph stage at a landscape scale 2) determine the biotic and abiotic factors regulating recruitment success in different ponds and 3) make predictions

about the importance of variations in recruitment success on the metapopulation structure and regional persistence of H. albopunctatus metapopulations.

Methods

Study Area and Site Selection

I studied breeding populations of H. albopunctatus from 1999-2002 at ephemeral ponds in the Shire of Kellerberrin, Western Australia, approximately 200 km east of Perth.

Audio-censuses to detect calling males were undertaken during the autumn (March-

May) breeding season of this species. A breeding population was defined as any site containing one or more calling male frogs that could represent a discrete water body when filled.

Once an initial population was located, I located all other populations of calling males within a two kilometre radius of this point. This design was then replicated in an area ten kilometers to the north of this initial cluster of populations. There were thus two spatially distinct clusters of breeding populations (Fig. 4.1).

Recruitment to Tadpoles

An earlier chapter established the average clutch size of H. albopunctatus of 391.2 ±

21.2 eggs/clutch, and there was no evidence of double-clutching or clutch splitting by females (Chapter 3). Egg mortality is also low in H. albopunctatus, averaging only

2.8% (Chapter 3). These data were used to estimate egg inputs into ponds given that I counted breeding burrows at all sites and only 37% of breeding burrows contained eggs

(Chapter 3).

Tadpole Abundance

From 2000-2002, 48 wetlands were sampled for tadpoles from the commencement of the first winter rains until tadpoles metamorphosed or ponds dried (usually July-

September). At each site that contained water, sampling was undertaken with a 1 mm mesh sweep net (350 x 250 mm opening, 50 cm deep). Smaller puddles were sampled with small hand-held dip nets. Given the small size of most ponds, they were sampled to exhaustion. Captured tadpoles were sorted in pond water and counted, identified, examined for the presence of limbs and then returned to the pond. This provided a measure of abundance and developmental stage.

Pond properties

Information on factors that could affect the abundance of tadpoles was collected prior to each sampling event to avoid disturbance resulting from sampling activities.

Conductivity and dissolved oxygen were measured with a hand held Yeocal® meter and maximum depth and pond area were measured with a tape measure. Rainfall was recorded from rain gauges near, or at, each population. Wetland hydroperiod was defined for each pond as the total number of days (not necessarily consecutive) that the wetland held water (Babbit and Tanner 2000).

Recruitment to metamorphosis

Successful recruitment was determined by the capture of metamorphs on their spring emergence from breeding ponds. Metamorphs were captured in a continuous pitfall trap and drift fence array surrounding breeding sites or were hand-captured during night spotlighting at each site. For sites at which no emergent metamorphs were captured,

successful recruitment was designated if water remained in ponds at the cessation of sampling, and large tadpoles with well-developed front and rear limbs (ie. Gosner stages

38 and above) were still present.

Statistical analyses

To further investigate the factors affecting recruitment to metamorphosis pond properties for those ponds that produced metamorphs and those that did not were compared with a Mann-Whitney U test.

I generated a correlation matrix to look at the relationship between pond variables measured above including maximum, minimum and average salinity, depth, area, tadpole number, tadpole density, egg numbers and dissolved oxygen (DO) If any variables were highly correlated ( r > 0.8) the less biologically relevant variable was removed. To reduce any other problems generated by correlation between variables the data were subjected to a principal components analysis (PCA without rotation) using the

PCA module of StatistiXl. Principal component scores for each pond were then used in a Mann-Whitney U Test to test for biotic and abiotic differences between ponds with and without metamorphs.

Results

Distribution of anthropogenic ponds

Intensive breeding site surveys located a total of 48 breeding ponds in 2000: 35 in the southern section and 13 in the northern area (Fig. 4.1).

Recruitment to tadpoles

Estimated annual egg inputs for each site varied from 281 to over 28 000 eggs per site

(Table 4.1). There was a notable reduction in clutch numbers in 2001 and 2002, due to drought. Recruitment to the tadpole stage from egg masses was very low with embryonic survival averaging just 1.25% (range 0.04-13.68%) excluding an unusually high survival rate of 65.75% for pond 3b in 2000 (Table 4.1). Of 48 ponds monitored from 2000-2002, on average, 86.11% of ponds failed to ever recruit tadpoles. In the best year of the study (2000) average larval recruitment success was only 27.08%, and in the worst year (2002) there was complete recruitment failure. Hydroperiod was the leading cause of recruitment failure with an average of 77% of all ponds failing to fill with water over the three year study (Table 4.2). Furthermore, only 7% of ponds held water for 60 days or longer (Table 4.2).

Tadpole abundance

Maximum tadpole counts were generally low and only three ponds recruited more than

300 tadpoles on a single occasion during the three-year study (Table 4.1; Fig. 4.2).

Where tadpoles were recruited, there was generally a rapid decline in abundance with time (Fig. 4.3). Tadpole abundance did not appear to be strongly related to hydroperiod, although greater numbers of tadpoles were generally found in ponds with a hydroperiod ranging from 50-60 days (Fig. 4.4). Salinity of ponds affected tadpole presence and abundance with no tadpoles being found in 5 breeding ponds with conductivities higher than 5000 µS (Fig. 4.5). The greatest number of tadpoles was present in ponds with conductivity measurements less than 1000 µS and in the range of 0-700 µS (Fig.

4.5).

Recruitment to metamorphosis

Survival to metamorphosis was low. Metamorphs were only recorded in 1999 and 2000, and not in 2001 or 2002. In the first year of the study (1999) only three ponds were monitored and recruitment data for these three ponds over the four years is presented in

Figure 4.6. As well as these sites, there was additional recruitment in 2000 at sites 3a (2 metamorphs), 4 (2 metamorphs) and 4a (3 metamorphs). Additionally, 15 and 71 tadpoles with well-developed front and rear limbs were still present at ponds 3a and 4 at the conclusion of sampling in 2000.

Statistical analyses

Only four of 23 sites containing water in 2000 produced metamorphs. Any analyses are therefore strongly biased by the unexpectedly small sample of ponds with metamorphs.

I restricted my analysis to making comparisons between sites with and without metamorphs using a Mann-Whitney U tests to minimise assumptions about variable properties. Table 4.3 presents the correlation matrix between all variables measured or derived from field measurements of pond properties. Because of their high correlations with other variables, minimum salinity, maximum depth and maximum area were eliminated from later analyses with maximum salinity, minimum and average depth, minimum area and minimum and maximum dissolved oxygen preferred because they represented the most biologically realistic and independent parameters (for example, the maximum pond area available is not as likely to be as important to survival as the minimum pond area). Additionally, maximum tadpoles and estimated egg input were included as biological variables for the same reasons. My goal was to generate an initial picture of what variables might affect metamorph success.

PCA analyses of both models are shown in Tables 4.4 and 4.5 and Fig. 4.7 and 4.8.

Generated component loadings were eliminated from further consideration if their

correlation with the component scores was not significant at the 0.01 level for the

Pearson correlation co-efficient (Norman and Streiner, 2000) a critical value of 0.575 for n = 19. For the abiotic model, average and minimum depth and hydroperiod explained most of the variance of PC1, however salinity was the only significant variable in PC2. In this model PC1, PC2 and PC3 accounted for 79.2% of the cumulative variance. For the full model including biotic and abiotic variables, PC1, PC2 and PC3 accounted for 65.8% of the total variance. Average depth, minimum depth and hydroperiod accounted for most of the variance in PC1 for the full model, with only minimum eggs and maximum salinity significant factors in PC2.

A Mann-Whitney U test between ponds with and without metamorphs based on casewise PCA scores of PC1 and PC2 revealed no significant differences between ponds (Abitoic model: PC1, MW U = 47, p = 0.1 ; PC2 MW U= 46,, p = 0.124; Full

Model, PC1, MW U = 47, p = 0.1 ; PC2 MW U= 47, p = 0.1 ) for either set of component scores. PC1 and PC2 scores are plotted by pond type (i.e. ponds with or without metamorphs) in figures 4.7 and 4.8.

Discussion

Recruitment Success to metamorphosis

I found that recruitment success to both the larval stage and to metamorphosis was generally poor and highly variable between ponds. Analysis of both biotic and abiotic factors did not give any clear cause of differences in success to metamorphosis .

Recruitment in amphibian populations is highly variable and driven primarily by fluctuations in larval survival (Alford and Richards 1999). The timing and length of

pond hydroperiod is often the most critical determinant of larval survival and subsequent recruitment success (Richter et al. 2003; Pechmann et al. 1989; Semlitsch

1987). If ponds dry before larvae have reached a minimum critical size for metamorphosis, they may desiccate and die (cf. Semlitsch 1983; Pechmann et al. 1989).

On average, 77% of the 48 ponds failed to fill with water during the three years of monitoring. Additionally, only 7% of ponds held water for more than 60 days. Lee

(1967) reported a larval duration of 90-150 days for all Heleioporus species and a larval duration of 128 days for a single H. albopunctatus egg clutch raised under laboratory conditions. However, of the four ponds that recruited metamorphs in 2000, two of these had hydroperiods of 50 and 55 days. Furthermore, pond 2 with a hydroperiod of just 50 days produced the largest number of metamorphs in the study (43 individuals). These conflicting results suggest the presence of other regulatory factors such as population- specific rapid larval growth, plastic growth patterns or tadpole density-dependent effects

(Wilbur and Collins 1973).

Intraspecific and interspecific density-dependent regulation of anuran larvae is well documented in the literature (eg. Smith 1983; Newman 1988; Wilbur and Fauth 1990a) and is an important feature of many amphibian breeding ponds. In the present study,

Mann-Whitney U Tests revealed that maximum tadpole number did not differ significantly between ponds with and without metamorphs. It is important to note that tadpole density may be an important factor in some years when conditions are optimum and recruitment is high. Blaustein and Margalit (1996) found that Bufo larvae were smaller and reached metamorphosis later at high densities and Reading and Clarke

(1999) reported that tadpole mortality in Bufo bufo was density-dependent, being proportionately higher when initial tadpole numbers were high.

Hydroperiod and Pond Features

Longer hydroperiods may favour the proliferation of predators (Smith 1983; Travis et al. 1985; Caldwell 1994, Tyler et al. 1998), and competitive interactions between species (Wilbur 1987; Wilbur and Fauth 1990b) both of which act to reduce recruitment success in temporary wetlands. Rowe and Dunson (1995) reported that survival rates of

Rana sylvatica larvae were much lower in temporary wetlands with a longer hydroperiod. Data on predation and competition are not available for this study.

Nonetheless, these biotic factors and their interactions with hydroperiod, may play an important role in regulating the recruitment success of populations.

Although there was no significant difference between ponds with and without metamorphs, the depth, hydroperiod and area of ponds contributed to most of the variance in PC1 of analyses. Further studies are required to investigate the importance of these factors on recruitment success.

Aquatic microhabitat may also have a significant influence on tadpole survival and subsequently, recruitment success. Several H. albopunctatus breeding ponds contained woody debris and submerged aquatic vegetation. This was not quantified, but other studies have found that aquatic microhabitat may influence recruitment and tadpole survival. Semlitsch and Reyer (1992) found that tadpoles in ponds containing predators spent more time in aquatic refuges (plastic aquarium plants) than those in predator-free treatments. Conversely, Phrynomantis tadpoles avoid submerged aquatic vegetation that may have harbored predatory freshwater turtles (Rodel 1999). Similarly, tadpoles of

Agalychnis callidryas also avoid bottom microhabitats to reduce their risk of predation from shrimps (Warkentin 1999). Pond microhabitat may be an important factor influencing survival of H. albopunctatus tadpoles, and requires further investigation.

Impacts of salinity

Mann-Whitney U tests showed no significant differences between ponds with and without metamorphs for PC2, heavily loaded by salinity in the abiotic model (Table 4.5) suggesting salinity was not a significant factor in determining success to metamorphosis if tadpoles were recruited. However, salinity may affect survival into the tadpole stage or determine whether any eggs are deposited at a pond. Although breeding ponds had conductivities of up to 21700 µS, tadpoles were not present in any waterbody with a conductivity exceeding 5000 µS (n = 5) and the majority of ponds wih tadpoles (13/18) had conductivities below 1000 µS. Salinisation is recognized as a major threat to frogs, particularly in the Australian environment (Ferraro and Burgin 1993) but there is a lack of published data on the salinity tolerances of Australian frogs (Hazell 2003). Main

(1990) highlighted salinisation as a probable cause of decline (based on museum specimens) for H. albopunctatus due to the loss of stream-line and valley breeding sites to salinisation. Baumgarten (1991) examined the salinity tolerances of H. albopunctatus tadpoles. Early-stage tadpoles experienced high mortality at approximately 8500 µS for the first 5 days, but mortality was constant for later stage tadpoles at 10 000 µS.

Baumgarten (1991) reported tadpole survival in conductivities of nearly 13 000 µS for

38 days. No tadpoles in this study were found in water with a conductivity exceeding

5000 µS, although one breeding site contained water with a conductivity of 21 700 µS.

This implies that although tadpoles in the laboratory have the ability to survive in relatively saline environments, they do not do so in a field situation although clearly further data is needed to establish this relationship for H. albopunctatus. Other factors may be regulating salinity tolerances. Baumgarten (1991) reported that adult H. albopunctatus showed a distinct preference for lower salinity soils. Active choice of less

saline breeding sites by adults may therefore play an equally important role in shaping the outcome of recruitment in a saline landscape.

Metapopulation persistence and management implications

While a number of studies have highlighted the importance of anthropogenically- modified landscapes for amphibian breeding and conservation (eg. Hecnar and

M’Closkey 1996; Stumpel and van der Voet 1998; Babbitt and Tanner 2000), Hazell

(2003) highlighted the lack of research into the impacts of habitat fragmentation and landscape change on Australian amphibians. Only a few studies have attempted to examine the impacts of anthropogenic change on populations (cf. Main 1990) or investigated historical causes of declines (eg. Wardell-Johnson and Roberts 1991). The historical impacts of agriculture on the West Australian environment are well documented in the scientific literature (eg. Hobbs and Hopkins 1990; Hobbs 1993;

Saunders 1989; Saunders et al. 1993) and numerous studies have examined the effects of habitat fragmentation and loss on amphibians.

The primary focus of this previous research has been the impacts of habitat fragmentation and genetic effects at the landscape level (e.g. Daly 1996, Hines et al.

1999; Hitchings and Beebee 1998; Roberts et al. 1999; Seppa and Laurilla 1999) and the influence of pond or habitat characteristics on regional populations (Vos and

Stumpel 1995; Vos and Chardon 1998; Baker and Halliday 1999; Hazell et al. 2001).

Few of these studies have attempted to examine population-level or demographic impacts of landscape change on amphibians. Hazell (2003) highlighted the need for such studies to examine either the stage of the lifecycle affected by these threats, or the scale of the threat (i.e. individual, population or regional scale).

In the context of this study, the primary impacts of habitat loss and fragmentation on the recruitment success H albopunctatus are at both the regional scale and the population level.

On a regional scale there is loss of natural wetland breeding sites to salinity, an enforced switch to anthropogenic breeding habitats (ditches and dams) and isolation of breeding sites by a matrix of cleared land. At the population level poor recruitment is due to pond desiccation.

Population Level

The three-year recruitment success of the 48 populations monitored was poor. Analyses indicated that although pond parameters did not have a significant impact on metamorphosis, hydroperiod was the most important determinant of recruitment success to metamorphosis indicating population regulation at a local scale, with regional implications. Larger ponds that hold water for longer have the highest and most consistent recruitment success. The majority of breeding sites had a very short hydroperiod or failed to fill with water.

The use of unfavourable breeding sites (ie. those with short hydroperiods, high salinities or small areas) may exacerbated by the breeding biology of this species. Adult

H. albopunctatus have no contact with water and presumably utilise cues such as soil moisture or landscape position when selecting breeding sites. A large number of less suitable sites are frequently used and the breeding site fidelity exhibited by females may enhance the problem as adults continue to breed in unfavourable ponds. Cues for egg laying that originated in a natural landscape, may not be a good indicator of pond quality in anthropogenic landscapes where differences in soil type, pond lining and predation pressures exert a greater and less predictable influence on recruitment success.

This situation is epitomized by the large number of frogs found breeding in damp areas of open paddocks where no surface water was observed. This leads to the concept of an ecological trap, with adult frogs selecting and returning annually to anthropogenic breeding sites with continual recruitment failure (e.g. Schlaepfer, 2003).

Continued recruitment failures will have a profound impact on the survival of local

H. albopunctatus populations, potentially resulting in local extinction as populations age or suffer inbreeding effects. However, populations are best viewed as dynamic entities existing as regional metapopulations (eg. Berven and Grudzien 1990; Sjogren

1994) rather than single isolated populations. The existence of metapopulations for this species may therefore be essential to ensure its long-term persistence.

Regional and Metapopulation Effects

Three-year mark-recapture population studies of 4 populations of H. albopunctatus revealed a low-moderate site fidelity (average 10.26%) with dispersal to neighbouring populations recorded at all sites (chapter 5). This provides some evidence for the existence of a metapopulation structure for this species at least based on adult data. The site fidelity and dispersal capacity of tadpoles and metamorphs are unknown, but since good rainfall is rare dispersal of these life stages must generally be limited.

On a regional scale, few breeding populations of this species are now located in natural wetlands as the majority of these have become saline (cf. Main 1990). In addition, this species is now generally breeding higher in the landscape as a consequence of the salinisation of low-lying drainages and valleys (A.R. Main, pers. comm.).

Management actions should recognize the potential metapopulation structure of H. albopunctatus breeding populations and focus on providing a number of suitable “core” breeding habitats with large, deep ponds and a long hydroperiod. Given the likely dispersal between populations, a few core sites with successful recruitment may be sufficient to provide dispersing individuals with the opportunity to found new populations and bolster ailing populations, akin to the source-sink model described by

Harrison (1991).

The provision or enhancement of breeding habitats in agricultural landscapes is a novel and contentious topic in Australia. Some authors described the advantageous aspects of farm dams in providing breeding sites for frogs (eg. Tyler and Watson 1998;

Hazell et al. 2001) while others reported deleterious effects (Brock and Jarman 2000).

Whatever the outcome, further research is required into the ability of human-modified landscapes to support frog species. It appears that the switch to anthropogenic breeding habitats is a successful adaptive strategy for H. albopunctatus, and with careful management it may ensure the regional persistence of this species.

Figure 4.1: Location of northern and southern tadpole study sites north of Kellerberrin, WA (see Chapter 2 for location of Kellerberrin), in relation to native vegetation remnants (shaded).

1000 2000 2001 900

800

700

600

500

400

300

200 Maximum Numbe of Tadpoles Numbe Maximum 100

0 5 1i 3i 3l 4i 4l 3j 4j 1f 2f 3f 4f 1x 1a 1c 1e 2x 2a 2c 2e 3x 3a 3c 3e 3k 4x 4a 4c 4e 4k 1b 1d 1g 1h 2b 2d 2g 3b 3d 3g 3h 4b 4d 4g 4h 4n 4o 4p 4m Pond

Figure 4.2: Maximum number of tadpoles present at each site sampled from 2000-2002. No tadpoles were present at any pond in 2002.

2000 1x 1d 1200 1e 2x 1000 2a 2b 2c 800 2d 3x 3a 600 3b 4x 400 4a

Maximum Number of Tadpoles 200

0 July August Early Late September October September Sampling Time

2001

1x 600 2x 2b 500 2g 4x 4a 400 4b

300

200

Maximum Number of Tadpoles of Number Maximum 100

0 August September Sampling Time

Figure 4.3: Number of tadpoles present on each sampling occasion, at each of 48 ponds monitored in 2000 (top) and 2001 (bottom). Ponds not shown did not contain water.

1200

1000

800

600

400 Number of Tadpoles

200

0 0 102030405060708090100 Hydroperiod (days)

Figure 4.4: Relationship between hydroperiod and tadpole abundance for 2000 and 2001.

1200

1000

800

600

400 Numberof Tadpoles Numberof

200

0 0 1000 2000 3000 4000 5000 6000 Salinity (uS)

Figure 4.5: Relationship between water conductivity and tadpole abundance for all ponds in which H. albopunctatus tadpoles were present in 2000 and 2001. Ponds with conductivites of up to 27 000 µs but no tadpoles were present above 5000 µs as depicted above.

250

1999 200 2000

150

100

Number ofMetamorphs 50

0 123 Pond

Figure 4.6: Number of metamorphs recruited at 3 ponds (see Table 4.1) monitored for a four-year period (zero recruitment in 2001 and 2002).

Principal Component Plot

-1

-2

-3 -5 0 5 PCA 1 (28.9%)

Figure 4.7: Principal Component plot of PC1 vs PC2 for the full model including abiotic and biotic variables influencing metamorphosis of 19 H. albopunctatus ponds monitored in 2000. Open Squares indicate those ponds for which there was successful metamorphosis.

Principal Component Plot

3.5

3.0

2.5

2.0

1.5

1.0

0.5

0.0

-0.5

-1.0

-1.5 -5 0 5 PCA 1 (38.1%)

Figure 4.8: Principal Component plot of PC1 vs PC2 for the abiotic variables influencing metamorphosis of 19 H. albopunctatus ponds monitored in 2000. Open squares indicate those ponds for which there was successful metamorphosis.

Table 4.1: Population size, estimated maximum and minimum egg input, mortality and recruitment success of H.albopunctatus for all ponds containing water, 2000-2002. Range values for metamorphs indicate the use of actual metamorphs captured as well as late stage (limbed) tadpoles captured.

Year Site No. of Estimated Egg Estimated Egg Maximum % Mortality No. of % Mortality late Burrows Input Minimum Input Maximum Number of Egg-early Metamorphs Tadpole- Tadpoles Tadpole Metamorph 2000 1 48 6753 18252 94 98.6-99.5 0 100 1d 5 703 1901 2 99.7-99.9 0 100 1e 2 281 760 1 99.6-99.9 0 100 2 51 737 19393 395 46.4-98.0 43 89.1 2a 11 1548 4183 5 99.7-99.9 0 100 2b 20 2814 7605 35 98.8-99.5 0 100 2c 2 281 760 6 97.9-99.2 0 100 2d 4 563 1521 1 99.8-99.9 0 100 3 40 5628 15210 6 99.9-100.0 0 100 3a 10 1407 3802 55 96.1-98.5 2-15 72.7-96.4 3b 4 563 1521 1000 34.2 0 100 4 74 10411 28138 73 99.3-99.7 2-71 2.7-97.3 4a 1 141 380 52 63.1-86.3 3 94.2 5 ------2001 1 25 3517 9506 170 95.2-98.2 0 100 2 22 3095 8365 240 92.2-97.1 0 100 2b 5 703 1901 250 64.4-86.8 0 100 2g 1 141 380 5 96.4-98.7 0 100 3 15 2110 5704 0 100 0 N/A 4 30 4221 11407 500 88.2-95.6 0 100 4a 3 422 1141 50 88.2-95.6 0 100

Table 4.1 (cont.).

4b 4 563 1521 100 82.2-93.4 0 100 5 20 2814 7605 0 100 0 N/A 2002 1 23 3236 8746 0 100 0 N/A 2 31 4362 11788 0 100 0 N/A 3 17 2392 6464 0 100 0 N/A 4 34 4783 12928 0 100 0 N/A 5 10 1407 3802 0 100 0 N/A

Table 4.2: Frequency of hydroperiods for 48 ponds monitored from 2000-2002.

Hydroperiod (days)

0 1-15 16-59 > 60

2000 62.5% 4.2% 14.6% 4.2%

2001 68.8% - 14.6% 16.7%

2002 100% - - -

Table 4.3: Correaltion matrix of measured pond parameters for 19 H. albopunctatus breeding ponds in 2000.

Hydro. Min. Eggs Max. Eggs Avg. Sal Min Sal Max Sal Min Depth Max Depth Avg. Depth Avg. Area Min Area Max Min. Eggs 1.00 Max. Eggs 0.87 1.00 Hydroperiod 0.10 0.11 1.00 Average Salinity -0.22 -0.26 -0.12 1.00 Minimum Salinity -0.21 -0.25 -0.17 0.98 1.00 Maximum Salinity -0.17 -0.21 -0.08 0.98 0.92 1.00 Minimum Depth -0.15 -0.14 0.24 0.33 0.19 0.43 1.00 Maximum Depth 0.30 0.55 0.61 -0.05 -0.10 0.01 0.48 1.00 Average Depth 0.21 0.43 0.55 0.06 0.01 0.12 0.65 0.96 1.00 Average Area 0.28 0.61 -0.17 -0.22 -0.21 -0.23 -0.15 0.32 0.27 1.00 Minimum Area 0.37 0.50 -0.30 -0.19 -0.18 -0.21 -0.18 0.01 -0.01 0.88 1.00 Maximum Area 0.24 0.61 -0.12 -0.23 -0.22 -0.23 -0.14 0.38 0.32 0.99 0.82 Average DO 0.00 -0.05 -0.35 0.25 0.35 0.13 -0.45 -0.38 -0.36 0.16 0.28 Minimum DO 0.02 -0.08 -0.30 0.36 0.42 0.28 -0.25 -0.29 -0.28 0.05 0.22 Maximum DO -0.02 -0.02 0.08 -0.03 0.07 -0.15 -0.61 -0.20 -0.30 0.15 0.20

Table 4.4: Results of PCA of pond parameters affecting metamorphosis for 19 ponds sampled in 2000 including biotic and abiotic variables in the model.

Explained Variance (Eigen values) Value PC 1 PC 2 PC 3 PC 4 PC 5 PC 6 PC 7 PC 8 Eigenvalue 2.315 1.626 1.327 1.228 0.635 0.502 0.224 0.143 % of Var. 28.932 20.325 16.586 15.351 7.941 6.274 2.800 1.792 Cum. % 28.932 49.257 65.843 81.193 89.134 95.408 98.208 100.000

Component Loadings Variable PC1 PC2 PC3 Hydroperiod 0.658 0.139 -0.430 Max Salinity 0.226 -0.717 0.480 Average Depth 0.843 0.306 0.279 Min Depth 0.803 -0.194 0.401 Min Area -0.389 0.565 0.507 Min DO -0.520 -0.280 0.497 Min Eggs -0.067 0.644 0.161 Max Tads 0.221 0.386 0.377

Table 4.5. Results of PCA of pond parameters affecting metamorphosis for 19 ponds sampled in 2000 including only abiotic variables in the model.

Explained Variance (Eigen values) Value PC 1 PC 2 PC 3 PC 4 PC 5 PC 6 Eigenvalue 2.287 1.412 1.052 0.732 0.320 0.197 % of Var. 38.112 23.536 17.534 12.207 5.328 3.282 Cum. % 38.112 61.648 79.182 91.389 96.718 100.000

Component Loadings Variable PC1 PC2 PC3 Hydroperiod 0.697 -0.351 -0.078 Max Salinity 0.259 0.886 -0.160 Average Depth 0.829 0.003 0.439 Min Depth 0.787 0.387 0.190 Min Area -0.406 -0.032 0.883 Min DO -0.512 0.594 0.114

Chapter 5: Population ecology of the frog Heleioporus albopunctatus in the central wheatbelt of Western Australia: Survival and population size of adults.

Introduction

Global interest has recently focused on the widespread and often catastrophic decline of amphibian populations. The disappearance of species from a number of isolated and pristine habitats has been well documented (Barinaga 1990; Fellers and Drost 1993;

Pounds and Crump 1994; Pechmann and Wake 1997). In the midst of such declines, there has been intense speculation on the cause and nature of these events and whether or not they are the product of human disturbance. It is often difficult to distinguish between declines resulting from human activities and those resulting from natural population fluctuations (Pechmann et al. 1991; Blaustein et al. 1994; Sarkar 1996) and

Alford and Richards (1999) cautioned on the importance of being able to separate these two events. To assess amphibian declines or construct even the most rudimentary population models, data on basic population parameters are required.

Detailed demographic studies on amphibians are scarce (even for those species not obviously in decline), and few have attempted to estimate population size and survival using detailed mark-recapture models (but see Gittins 1983; Ramirez et al. 1998; Wood et al. 1998; Conroy 2001; Gillespie 2001; Marunouchi et al. 2002). Many studies on amphibian population dynamics have been restricted to one population or site (eg.

Richter et al. 2003) or a short time period, and are consequently of limited use in investigating regional changes or the long-term persistence of species.

Alford and Richards (1999) constructed a null model for amphibian populations in which they argued that most amphibian populations should decrease more often than

they increase due to highly variable recruitment and less variable adult mortality. The model included highly variable recruitment between years and between ponds and with egg and tadpole mortality varying by several orders of magnitude, but relatively constant survival in terrestrial adult stages (Alford and Richards 1999).

Knowledge of the inherent variance in the demographic parameters of non-declining species will set a standard against which to compare the population dynamics of species suspected to be declining (Alford and Richards 1999). I report here on a three year mark-recapture study of five breeding populations of Heleioporus albopunctatus to quantify the variance in demographic parameters of a non-declining frog species in a highly human-modified agricultural landscape. I was interested in estimating the probability that a frog captured during a breeding season would survive until the next breeding season and in the recapture probability of individuals between years. I also wanted to test the hypothesis that survival between populations was sex-biased.

Methods

Selection of Study Sites

Five breeding populations of H. albopunctatus were selected in 1999 near Kellerberrin in the semi-arid, agricultural region of south-western Australia (the “wheatbelt”; cf.

Saunders et al., 1993). Each site contained a minimum of 40 calling males at the time of selection.

Three breeding populations (Wilkins, Morley and Morley West) were located between 12 and 14 km north of Kellerberrin (Fig. 5.1). The other two breeding populations (Leake and Hammond) were located approximately 25 km north of

Kellerberrin (Fig. 5.1).

This design ensured that there were two spatially distinct clusters of breeding populations, separated from each other by a distance of ten kilometers which I assumed was beyond the dispersal distance of adults of this species. This design allowed for two independent replicates assuming that frogs might move between close sites.

Breeding populations were all sited in intensively human-modified environments and comprised non-natural breeding ponds. The populations Morley (31o30’37”S

117o43’52”E) and Wilkins (31o30’40”S 117o43’51”E) were both located on private property, in and around historical (1950s) sand-mining areas that comprised large, 2m deep depressions from which sand had been extracted. The upland vegetation surrounding the depressions was composed of Banksia prionotes woodland with a mixed herbaceous understorey. The area surrounding the vegetation remnants was arable farmland. The depression in which frogs bred was largely devoid of vegetation.

Morley West (31o30’55”S 117o43’23”E) was situated on private property in a clay- lined “salinity interceptor bank”, approximately 20 m long and 2.5 m deep. It had no native vegetation and was surrounded by arable pastureland, with no upland vegetation surrounding the breeding site.

Leake (31o25’38”S 117o44’57”E) was a roadside ditch on Crown land, approximately 10 x 20 m. It was inside a small corridor of native Melaleuca and Acacia species, in an otherwise open cropland with very high weed invasion by wild oats and other grasses. Hammond (31o25’43S 117o44’49”E) was located on private property to the west of Leake and was a damp patch of pasture, with no native vegetation present, completely surrounded by fields.

Trapping Methods

Capture and marking of frogs at breeding populations was undertaken from 2000 –

2002. H. albopunctatus migrates to breeding sites on the first autumn rains, usually in

March/April (Lee 1967). Trapping commenced with the start of the breeding season and ceased when all frogs had migrated from breeding sites. Migrating frogs were intercepted with drift fences and pitfall traps.

All breeding sites were encircled with a 30 cm high, aluminium mesh, drift fence.

Pitfall traps consisted of twenty litre white plastic buckets installed flush with the soil surface. Buckets were placed at approximately 30 m intervals, comprising even numbers of buckets inside the fence and outside the fence. Traps were checked each evening and every morning to remove captured frogs. Time was also spent each night searching the fence to hand capture any frogs that were migrating in or out of the breeding site.

Marking and Sexing

Captured frogs were identified with a unique mark by toe-clipping, using the scheme in

Hero (1989) which minimizes the number of toes removed.

Frogs could be sexed during the breeding season by the presence of black nuptial spines on the thumbs of males (Lee 1967). Individuals were also examined for the presence of eggs, which were easily discernible in gravid females.

Survival Rate Estimation

To estimate the probability of survival and recapture, I used Program MARK (White and Burnham 1999). MARK was used to estimate differences in the rates of survival

and recapture between and within populations, over time and between sexes, using the

Cormack-Jolly-Seber (CJS) model. Model notation followed Lebreton et al. (1992) as follows:

Øi = survival probability from time i to i+1 pi = probability of capture or recapture at time i. g = sex(group)dependent survival t = time-dependent survival g*t = both sex(group)-dependent and time-dependent survival

The absence of subscript notation implies constancy ie. no specific co-variables such as time or group (eg. Øt or Øg) were present.

As well as examining population-based differences in survival and recapture probabilities, I was particularly interested in testing whether survival and recapture probabilities were sex-biased. Population studies were based on breeding aggregations that lasted only six weeks, and did not provide sufficient capture-recapture data to warrant analysis of survival within breeding seasons.

To estimate between-year survival and recapture rates, I used individual capture histories, pooling all data for captures of an individual during the breeding season.

Capture histories for males and females over three breeding seasons were constructed with the exception of Hammond, which was only studied for a two-year period.

Models were constructed in MARK to examine survival and recruitment. I started with the most inclusive models and tested the goodness of fit of the Cormack-Jolly-

Seber model Øg*t pt which included population and time-dependent survival and time- dependent recapture probability. Goodness of fit was tested with the bootstrap GOF program in MARK, for 100 iterations. Probabilities were calculated based upon deviance ranks and mean deviance and variance inflation factor (c-hat) from GOF

simulations were used to calculate new c-hat values for the model. As a conservative estimate, the highest c-hat value was chosen and the data adjusted accordingly. C-hat is a measure of the over-dispersion or under-dispersion of the data, with a value of 1 considered a perfect fit, >1 indicating over-dispersion and <1 indicating under- dispersion (Burnham and Anderson 1998).

Several candidate models were run and the most parsimonious model was chosen based upon its Akaike’s Information Criterion (AIC) or quasi-likelihood AIC (QAIC) rank and weight (Burnham and Anderson 1992). The QAIC is an AIC parameter resulting from c-hat adjusted data (Anderson et al. 1994).

Where model weightings were similar, model averaging was used to provide an average weighted estimate of survival and recapture parameters, based on all models in the candidate set (Buckland et al. 1997).

Population Size Estimation

Several estimates of population size were used. A conservative “minimum number known to be alive” estimate (“census population size”) was based on the actual number of field captures each year. Abundance estimates based on the Jolly-Seber model were calculated for each population in which recapture data were sufficient for analysis.

Estimates were only calculated using data for between-years captures.

Although it was technically possible to use within-years data, with each trip as a capture period, frogs were not resident for the duration of the breeding season and there was a continual immigration and emigration of breeding individuals. The treatment of each breeding season (one per year) as a capture-recapture period minimized these factors and ensured that all returning individuals had the same chance of being recaptured.

Jolly-Seber (J-S) population estimates could only be calculated for the second year,

2001, of the study at each population as estimates require n-1 and n+1 years of data to estimate n (Seber 1982). Population sizes could not be estimated for Hammond as there were no recaptures in the second year.

Sex Ratios

Sex ratios were calculated for each year, on the basis of the census population data.

Results

A summary table of the most parsimonious models for each population is shown in

Table 5.1. Survival and recapture estimates based on the most parsimonious models for each population, are displayed in Figures 5.2 and 5.3.

Bootstrap goodness of fit tests gave support to the fully paramaterised CJS model Øg*t pt, as an initial model for all populations (Table 5.1). The fully parameterized model was well supported for Morley (p = 0.45). The most parsimonious model for Morley was phi(.)p(.) based on AIC weighting (0.34), indicating constant survival (phi) and constant recapture (p). Both recapture and survival rates for this population were low

(Fig. 5.2).

For the neighbouring Wilkins, the fully paramaterised model had good support (p =

0.66). For this population phi(.)p(.) was the most parsimonious model with an AIC weighting of 0.34.

The fully paramaterised model had a moderate fit for Leake (p = 0.13). The most parsimonious model was also the fully reduced model phi(.)p(.) with a strong AIC weighting (0.44), indicating constant survival and recapture (Figure 5.2).

The fully paramaterised model had a good fit for Morley West (p = 0.26). The most parsimonious model for this population was phi(g)p(.) representing sex-dependent survival and constant recapture (AIC weighting = 0.44). Female survival was half that of male survival and recapture rates were moderate.

Survival of males and females between years could not be compared for Hammond due to unequal representation of sexes. This site was only studied for two years as it was discovered late in the study. During the first year only females were captured, while during the second year only males were captured. This site was excluded from further analysis.

The survival probability (phi) was estimated as 1 for Leake i.e. 100% survival of males and females between years (Fig. 5.2). Estimated survival ranged between 0.335 and 0.696 for the remaining populations. A sex-biased survival trend was apparent for

Morley West, with female survival half that of males (Fig. 5.3), despite recapture rate being constant for both sexes.

Census and J-S population size for 2001 differed greatly and annual census breeding population sizes fluctuated greatly between years (Table 5.2). Data for all populations with three years of data (except for Leake), revealed that census population size was greatest in 2000 and lowest in 2001.

Sex ratios also fluctuated greatly between years, ranging from slightly male biased in

2000, to greatly female biased in 2001, for all populations except Morley West in 2001.

All populations were again male biased in 2002. Although population size remained similar, Hammond (with only two years data) swung from completely female in 2001 to completely male in 2002.

Annual census breeding population size was strongly correlated with rainfall during the March to May breeding season. Hammond (r2 = 1.00), Wilkins (r2 = 0.95) and

Morley West (r2 = 0.85) all showed a strong correlation with rainfall, with capture-rates increasing in years of higher March-May rainfall. Leake (r2 = 0.53) and Morley (r2 =

0.57) showed a weaker relationship.

Discussion

Survival and recapture estimates

The most parsimonious model for all sites except Morley West, recognized constant survival and recapture and no sex-biased differences within breeding populations

(Figure 5.2). In contrast, Morley West was characterised by lower female survival

(Figure 5.2). Most studies on amphibian survival (usually on breeding populations as for this study) have reported no sex-based differences in annual survival. Williamson and Bull (1996) noted no differences in survival of male and female Crinia signifera, a non-burrowing Australian species.

Conversely, Lemckert and Shine (1993) reported higher female survival in Crinia signifera and Licht (1974) reported higher female survival in Rana pretisoa, an aquatic breeding species. There are few comparisons available for other burrowing species.

Jehle et al. (1995) observed no significant sex-based difference in survival of the burrowing frog Pelobates fuscus. A summary of the literature on anuran survival is presented in Table 5.3, with the survival estimates for H. albopunctatus inserted for comparison. Only 25% of the studies reported a sex-biased difference in annual survival

(Table 5.3).

There is no evidence of differential male and female survival at most sites and both sexes have a good chance of surviving to the next breeding season. Survival does not decline with time since initial capture. The reasons for lower female survival at Morley

West are unclear. This population was situated in a relatively isolated location, surrounded by a matrix of open fields and was at least one kilometre from the nearest neighbouring vegetation. This contrasted with the remaining populations all of which

(apart from Hammond) were well connected to each other or other vegetation remnants by roadside corridors, and were surrounded by upland native vegetation.

Studies on the congeneric burrowing frog H. eyrei (Bamford 1992) have shown that this species uses upland habitat during the non-breeding season, although this has not been established for H. albopunctatus. Lee (1967) suggested that Heleioporus are nomadic outside the breeding season, but described the use of upland habitat adjacent to a swamp on Rottnest Island by H. eyrei. Lee states that individuals were never observed in the actual swamp during winter or summer but documented their use of the upland vegetation. Lee also described the use of summer burrows by H. albopunctatus.

If riparian vegetation is important non-breeding habitat for H. albopunctatus, the absence of suitable areas at Morley West may have resulted in obligate dispersal of males and females. Gravid females may be more susceptible to predation by foxes and birds and this may have resulted in reduced survival of females in this population.

Alternately, the dispersal capacity of males across cleared land may be greater and this may result in more males being present from year to year. Further movement studies are clearly required to determine habitat use and the nature of male and female dispersal in the agricultural matrix.

Contrary to the female-biased mortality reported here for one site, Lemckert and

Shine (1993) reported increased mortality of male C. signifera at breeding sites due to their conspicuousness during calling, and the relatively cryptic nature of females by comparison. Few studies on anuran survival have examined more than one population

and comparative literature on differential population survival is scarce (but see Berven

1982; 1990).

Survival rates for H. albopunctatus are within the range 0.335 to 1 and averaged 0.59 across all populations (Figs. 5.2, 5.3). This compares favourably with reported survival rates in the literature (Table 5.3). Lowest survival estimates for H. albopunctatus (0.21

Table 5.2) are still high in comparison to the estimates of 0.04-0.24 for Rana sylvatica

(Berven 1990) and 0.02-0.05 for R. erythraea (Brown and Alcala 1970), both species studied in continuous, natural habitats. The lowest survival estimates for H. albopunctatus populations in fragmented landscapes, are generally well within the range expected from other anuran population studies (Table 5.3).

The high survival and low recapture rates for Wilkins, Leake and to a lesser extent,

Morley, are due to most recaptures of marked frogs occurring in the last year of the study. It is possible that H. albopunctatus does not breed every year. Despite Main’s

(1968) claim that breeding in Heleioporus is strictly seasonal and independent of rainfall, H. albopunctatus may not breed during years of drought and this could explain the low numbers in 2001. In this year rainfall was below average due to severe drought conditions in south-west WA (Bureau of Meteorology 2001).

Population Size

Jolly-Seber population size estimates varied between sites. Morley had the largest estimated population size (86). Morley West had the smallest estimated population size

(39). Estimated population sizes indicate that H. albopunctatus populations are generally small. There are few comparable population studies on burrowing frogs in

WA. Studies of H. eyrei at Rottnest Island gave a census population size of 280 adult

frogs (Lee 1967). Population size estimates for Neobatrachus pelobatoides near Perth were placed at 116 (Main 1968).

Population estimates differed substantially from census population sizes. A direct comparison was not available as J-S population sizes were only estimable for 2001.

Estimates for Leake were twice that of 2001 census size, and four times 2001 census size for Morley. Estimates for Morley West were very similar to census size.

Variations in inter-annual population size may reflect differences in rainfall between years. The lowest annual population size at all populations apart from Leake, occurred in 2001, a year of drought. The strong relationship between rainfall and population size indicates the importance of climatic variables in regulating breeding site attendance and may indicate the presence of temporary emigration in response to dry years. Henle

(2001) argued that the occurrence of temporary emigration in amphibian populations will bias parameter estimates. Schmidt et al. (2002) responded that the correct construction of capture histories with the use of GOF tests and appropriate models as used here, will detect violations of CJS assumptions and correct parameter estimates will be obtained.

Few comparable long-term population studies have been undertaken on Australian frogs (Littlejohn et al. 1993). Humphries (1979) noted large fluctuations in population size related to rainfall and Osborne (1989) attributed declines in Pseudoprhyne coroboree to severe drought. Population sizes of Crinia insignifera on Rottnest Island varied between 261 and 1341 over a ten-year period (Main 1968). Main (1968) found that population size was closely related to rainfall with the lowest estimates being from years of low rainfall and vice versa.

The significant influence of rainfall on amphibian breeding activity has been widely documented in the literature (eg. Friedl and Klump 1997; DiMauro and Hunter 2002;

Jensen et al. 2003 ) although a few studies have found no relationship (Meyer et al.

1998).

Sex Ratios

Overall observed sex ratios were generally 1:1 with a slight female bias (Table 5.2).

Annual sex ratios also reflected this trend with the exception of 2001, the drought year.

In this year, sex ratios were female biased for all populations, apart from Morley West which was not significantly different from other years (Table 5.2). Reported sex ratios in anurans include 1:0.59 for Pelobates fuscus (Hels 2002), 1:1.78 and 1:2.64 for two consecutive breeding years of Rana pretiosa (Licht 1974), 1:0.47 – 1:1.07 for a 6 year study of P. fuscus (Jehle et al. 1995) and 1:1.02-1:12.3 for R. sylvatica (Berven 1990).

Lee (1967) documented an overall sex ratio of 1:0.39 males:females for H. eyrei breeding on Rottnest Island. This sex ratio is much lower than the data presented for H. albopunctatus a difference that may be due to the population dynamics of an island population (eg. a genuine shortage of females or lower female survival). No equivalent published data exists for mainland populations of H. eyrei or other Heleioporus species.

In summary, it is apparent that sex ratios are variable and inconsistent between both species and years. This may be closely linked to rainfall and other undetermined factors that affect breeding conditions.

Survival is a core parameter in population dynamics and determining adult survival of H. albopunctatus populations is a key step towards understanding the autecology of this species. Alford and Richards (1999) argued that an understanding of the autecology of amphibians and the ecology of metapopulations is an essential requirement in understanding amphibian declines, and these data are often lacking in amphibian studies.

As predicted by Alford and Richards (1999) null model of amphibian population behaviour, adult survival in this study was generally high and constant compared to juvenile stages (Chapter 4). The most parsimonious population model for all four populations was one that included a constant survival term. Populations differed in the magnitude of adult survival, with a range of 0.2-1 for the five populations. Survival and population size were strongly positively correlated with rainfall and the nature of this association varied between populations. The implications of these findings for H. albopunctatus, a non-declining frog species, are that variance in adult survival is low or constant but populations clearly differ in adult survival with some having a higher mortality (eg. Morley). This implies that populations may need to be managed on an individual basis with regard to their regional persistence.

Sex-ratios as well as the degree of response to rainfall are also important factors affecting population demographics. Alford and Richards (1999) highlighted the importance of larval and egg mortality in regulating recruitment success, and further studies are required to investigate the role of these factors in regulating population size.

The weakness of this study like many other amphibian populations studies described by

Alford and Richards (1999) is that only a three year dataset is available for analysis.

Long-term monitoring is required to elucidate the importance of seasonal effects such as rainfall on breeding success and population size. The strength of this study, however, is that it focuses concurrently on 5 populations from two independent regions, and reduces the influence of stochastic events that may have profound effects on single populations.

Table 5.1: The most parsimonious models for four populations of H. albopunctatus. Model notation is described in the methods. Phi = survival probability, p = recapture probability and (.) = constant.

Population Most Parsimonious Model AIC weight

Morley phi(.)p(.) 0.34

Wilkins phi(.)p(.) 0.34

Morley West phi(g)p(.) 0.44

Leake phi(.)p(.) 0.44

Table 5.2: Annual census population size, Jolly-Seber population estimates and sex ratios.

Census population Sex ratio (M:F) Population J-S Average 2000 2001 2002 2000 2001 2002 population annual sex

ratio

Wilkins 35 10 26 - 1:0.67 1:4 1:0.73 1:1.25

Morley 65 22 31 86 1:0.67 1:2.14 1:0.72 1:1.04

Morley West 56 41 64 39 1:0.75 1:0.86 1:0.6 1:0.72

Hammond - 12 8 - - 0:12 1:0 1:1.09

Leake 35 33 25 69 1:0.94 1:2 1:0.25 1:1.05

Table 5.3. Documented annual survival rate ranges for anurans. Numbers in brackets are midpoints of survival ranges. Species marked with an asterix (*) are those for which a statistically significant difference in male and female survivorship was reported. ** indicates a species in which there was a significant difference in survival between males and females in at least one year of the study. After Conroy (2001).

Species Sex Annual adult Reference

survival

Bufo bufo* M 0.47-0.56 (0.515) Gittins 1983

F 0.31-0.49 (0.4) Gittins 1983

B. calamita M 0.4-0.42 (0.41) Sinsch and Seidel 1995

Crinia signifera* M 0.32 Lemckert and Shine 1993

F 0.41 Lemckert and Shine 1993

C. signifera M 0.11-0.75 (0.43) Williamson and Bull 1996

F 0.16-0.76 (0.46) Williamson and Bull 1996

Geocrinia alba M 0.08-0.73 (0.405) Conroy 2001

G. vitellina M 0.13-0.61 (0.37) Conroy 2001

Heleioporus albopunctatus* M,F 0.201-1 (0.60) This Study

Pelobates fuscus M 0.13-0.50 (0.315) Jehle et al. 1995

F 0.19-0.38 (0.285) Jehle et al. 1995

Pseudacris nigrita* M 0.15-0.34 (0.245) Caldwell 1987

F 0.27-0.55 (0.41) Caldwell 1987

Table 5.3 (cont.).

P. ornate* M 0.25-0.73 (0.49) Caldwell 1987 F 0.43-0.74 (0.585) Caldwell 1987 Rana aurora M,F 0.69 Licht 1974 R. cascadae M 0.59 Briggs and Storm 1970 F 0.46 Briggs and Storm 1970 R. erythraea M,F 0.02-0.05 (0.0315) Brown and Alcala 1970 R. esculenta M,F 0.53-0.7 (0.615) Holenweg Peter 2001 R. lessonae M 0.21-0.45 (0.33) Sjogren 1991 R. lessonae F 0.11-0.41 (0.26) Sjogren 1991 R. lessonae M,F 0.72-0.84 (0.78) Holenweg Peter 2001 R. grylio M 0-1 (0.5) Wood et al. 1998 F 0-0.61 (0.305) Wood et al. 1998 R. pretiosa* M 0.45 Licht 1974 F 0.67 Licht 1974 R. sevosa** M,F 0.16-0.22 (0.19) Richter and Seigel 2002 R. sylvatica M,F 0.04-0.24 (0.14) Berven 1990 R. temporaria M,F 0.36 Loman 1984 R. temporaria M 0.16-0.51 (0.335) Elmberg 1990 R. temporaria F 0.05-0.33 (0.19) Elmberg 1990 R. vaillanti M,F 0.68 Ramirez et al. 1998

⇑ N

1: 15 000 000 1:125 000

Figure 5.1: Location of Heleioporus albopunctatus population study sites and major drainage systems, near Kellerberrin in Western Australia (see inset). The extent of the WA wheatbelt region is indicated by dotted lines in the inset.

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Morley 1.9 Wilkins Leake

1.4

Rate 0.9

0.4

-0.1 Survival Recapture

Figure 5.2: Annual survival and recapture estimates for the three populations with constant survival (phi) and recapture (p) ie. phi(.)p(.). Standard error bars are shown.

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1 Survival 2001 2002 0.9 Recapture 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0 MFMF M F MF

Figure 5.3: Survival and recapture estimates for Morley West, based on model averaging of the most parsimonious models phi(g)p(.) and phi(g)p(t). Error bars are shown. M = male and F = female.

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Chapter 6: Persistence and extinction risk of the Western Spotted Frog

Heleioporus albopunctatus in a fragmented agricultural landscape: a population viability analysis.

Introduction

Although H. albopunctatus currently appears secure (Roberts et al. 1999) detailed studies of genetic structuring, adult, larval and egg survival and dispersal indicate the presence of some small, inbred populations, many populations with high larval mortality, no or very low recruitment and moderate adult survival as well as a small number that have more positive demographic features (Chapters 2-5). This indicates the possibility of a source-sink metapopulation structure (see Harrison 1991) for H. albopunctatus breeding populations which would place importance on maintaining

“core” breeding habitats with successful annual recruitment to maintain this species at a landscape scale (Chapter 2).

Population viability analysis (PVA) is an effective, transparent and widely used process for determining the long-term impacts of stochastic and deterministic factors on the persistence of populations (eg. Brook et al. 1997; Reed et al. 2003). There has been contention about the accuracy and role of PVA in predicting extinction probabilities, particularly with scant data (e.g. Coulson et al. 2001). However, comprehensive recent studies have unequivocally advocated the importance and robust nature of PVA in producing unbiased predictions (Akcakaya and Sjogren-Gulve 2000; Brook et al. 2000).

Brook et al. (2002) further added that PVA is still the only universally accepted, transparent process for estimating extinction probabilities and that no suitable alternatives are currently available.

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The objective of this chapter was to utilise life-history data, largely collected in previous chapters, to model the long-term persistence of H. albopunctatus populations with the use of PVA. In particular I was interested in how various models of both local populations and metapopulations, responded to variation in demographic parameters and potentially catastrophic climatic events.

Methods

Program Used for PVA

I used Vortex Version 9.42 (Lacy et al. 1995) for PVA. This program is widely used for modeling wildlife populations (Lindenmayer and Lacy 1995; Brook et al. 1997; Towns et al. 2003) and can simulate the effects of deterministic forces, genetic, environmental and stochastic effects on wildlife populations (Lacy 1993). Inbreeding was not modeled, as this is unlikely to be a major impact in this system (Chapter 2).

Estimation of life-history traits

Key life-history data (Table 6.1) for PVA were mostly derived from population studies on H. albopunctatus in the central wheatbelt from 2000-2003 (Chapters 2-4). These data formed the basis for PVA input parameters. Further parameters for which there was no species-specific information (eg. juvenile survival) were estimated from the literature on anuran populations. Field studies and literature searches were used to estimate other

PVA input parameters as follow:

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Age: Lee (1967) stated that all Heleioporus species attain sexual maturity and breed only after two years of age. The upper age limit for H. albopunctatus in this PVA is derived in part from an unpublished honours thesis (Thom 1995) that estimated ages of up to 7 years for this species based on skeletochronology. Due to the small sample size

(n = 27) of this skeletochronology study and its lack of calibration, I conservatively used ten years as an upper age limit. A stable age distribution was not assumed as adults aged 0-1 were not present in the breeding population.

Reproduction: There is no evidence to suggest that H. albopunctatus is polygamous

(Chapter 3) and a 1:1 sex ratio and full availability of both sexes in the breeding pool was assumed. All females captured in each population were assumed to be breeding or have just bred as very few frogs have ever been sighted at breeding sites outside the breeding season. The average number of offspring per female was estimated as the average number of metamorphs successfully recruited per single egg clutch. This was a best-case scenario estimate derived from data obtained during studies of larval survivorship (Chapter 4) and embryonic mortality (Chapter 3). Final estimates were derived using the following formula:

Number surviving to metamophosis = (average clutch size) X (larval survival) X

(embryonic survival)

Survival: Mortality estimates for adults were derived from mark-recapture studies

(Chapter 5). Because I did not recapture any marked juveniles, I was unable to estimate juvenile survival to breeding age. I therefore used an average estimate of juvenile

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survival rates reported in the literature (based on a literature review and data presented in Conroy 2001, p52).

Population Size: Initial population size was based on the average population size estimated from five populations at which intensive mark-recapture was undertaken

(Chapter 5). Carrying capacity was an estimate of the largest potential population size taking into account the size of habitats and resource availability but this is realistically a best guess. Carrying capacity and population size varied between models as described below.

Constructing Basic Models

Since genetic and mark-recapture studies both suggested dispersal between populations

(Chapter 2, 5), a metapopulation model (rather than a single population) was used to represent the spatial structure of all populations in a landscape context, and investigate the impacts of habitat fragmentation. Three different models were constructed. All base scenarios used the parameters outlined in Table 6.1, but each model varied in initial population size, carrying capacity and dispersal (Table 6.2).

1. Generalized metapopulation model: 25 populations with an initial population size of

87, carrying capacity of 300 and dispersal rate 1.67% (calculated from an average of

dispersal between sites during mark-recapture studies). This represents an

optimistic best-case scenario for populations, based on averages of parameters

estimated from field studies.

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2. Population specific model: 25 populations with an initial population size varying

from 2-86 and a carrying capacity varying from 4-300, dispersal at 1.67%. For this

model, carrying capacity and initial population size varied with population and were

based on monitoring data (Chapter 4) of maximum burrow counts at each site as it

has been previously established that burrow counts provide a reliable estimate of

population size (Chapter 3). This represents a realistic model of populations

generating recruits at a broad landscape scale, by the inclusion of variability in

population size and carrying capacity, thus some populations are larger and often

more successful than others. This model does not include sites with zero recruitment

and thus represents the integrated set of “source” populations in a source-sink

metapopulation model (e.g. Harrison 1991).

3. True metapopulation: 5 populations with an initial population size varying from 12-

86, dispersal varied from 0-20% but carrying capacity was set optimistically at 300.

This reflects the actual known demographic values of sites actively monitored for

population size and survival estimates (Chapter 5) except for the estimated carrying

capacity which was based on a best guess estimate of the maximum number of frogs

that could be supported by each site (based on availability of breeding substrates,

proximity to water etc.).

For all models, the catastrophic effects of drought were based on long term (1962-2002) rainfall records collected weekly by landholders in the study area. In 2001 there was a drought resulting in reduced population sizes or attendance at breeding sites (Chapter

5). Using rainfall data for 2001 (averaged 39 mm for the March-May breeding period across 5 sites) and the long-term rainfall records, I calculated a drought (rainfall in

March-May ≤ 39mm) frequency of approximately one year in three (0.33). Captures of adults during drought years (Chapter 5) were used to estimate the impact of drought on

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population size and reproductive success. The impact of high rainfall (> 100 mm in breeding period and more than 400 mm p.a.) years on populations was calculated from rainfall records as one year in 3.5 (frequency of 0.275). Thus drought years were almost as frequent as good years for reproduction. The positive impacts of good years on recruitment were estimated as being of the same order of magnitude as the negative impacts of drought on populations.

I projected all scenarios for 100 years and one thousand iterations of each model were run to ensure stabilisation of the results (Brook et al. 1997).

Scenario Projections

Once I constructed the basic models described above, I wanted to investigate the response of metapopulations to a variety of demographic and environmental scenarios that were likely to impact populations (Table 6.2).

Pond Drying: the most obvious cause of recruitment failure from field studies

(Chapter 4) was pond drying. Based on field studies of mean larval survival during drought years (Chapter 4), I estimated that offspring production was reduced from 37 to

16.7 per female (a 55% reduction). This represented a worst-case scenario for recruitment.

Adult Survival: Two important influences on population size were juvenile and adult survival. Although no field data were available for juvenile survival due to zero recapture of 500 metamorphs marked, a mean literature estimate of 22.5% was used.

Worst-case scenarios for juvenile survival were modeled by increasing metamorph mortality to 90% and 95%. Similarly, the sensitivity of the models to adult mortality was investigated by increasing adult mortality to 90% and 95%.

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Environmental Variation: Climatic and environmental influences on populations are particularly important in areas of low and unpredictable rainfall such as occurs in the range of H. albopunctatus. The basic model assumed that all populations were equally impacted by weather conditions (i.e. EV = 1) and used drought and rainfall frequencies based on long-term rainfall data. To investigate the impact of environmental variation, two further scenarios were constructed. One scenario used an adjusted EV of 0.5. This is a realistic assumption based on the probability of local variation in rainfall occurring: e.g. in localized thunderstorms, and the persistently high recruitment success of some populations over others (Chapter 4). The second climate-related scenario comprised a reduction in the frequency of high rainfall years from 0.275 to 0.1, simulating a possible run of low rainfall years.

Dispersal: The importance of dispersal to metapopulation persistence was investigated by increasing dispersal by a factor of ten from the basic model.

Statistical Analysis

A two-tailed t-test was used to investigate the significance of differences for each scenario between the generalized, population-specific and “true” metapopulation models.

Results

Basic Models

All three models revealed that the metapopulation was robust with strong, continued survival and persistence over 100 years (Table 6.3). For all models, the growth rate of

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the metapopulation and local populations, was positive and regulated primarily by populations reaching and persisting at or near their carrying capacity. Final projected metapopulation size was greatest for the generalised model primarily due to the artificially high carrying capacity (300) assigned to all local populations (Table 6.2).

The population-specific model had a much stronger growth rate for both local and metapopulations although the probability of extinction (Pe) of 0.062 for local populations, was slightly higher than the Pe of 0 for the generalized model (Table 6.3).

The True Metapopulation model was also robust with a metapopulation size of

1453.11 and Pe of 0 and a mean local population size of 292.71 and Pe of 0.007 (Table

6.3).

Scenario Projections

The three models run were generally robust to changes in most demographic parameters, with metapopulations consistently more robust than local populations

(Table 6.3). For the generalised and true models, none of the projected scenarios resulted in an extinction of the metapopulation (Table 6.3). The population-specific model was also generally robust, but was particularly sensitive to variations in juvenile survival. Juvenile mortality of 95% resulted in a Pe of 1 for both local populations and the metapopulation. Conversely, 90% juvenile mortality still lead to good metapopulation (Pe 0.077) and local population (Pe 0.14) survival. The cutoff value for juvenile survival thus appears to lie between 90 and 95%. Similar variations in parameters such as adult survival did not significantly affect the Pe under the population specific model.

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Overall, extinction probabilities were low or close to zero for all scenarios, with the exception of a high local population extinction rate (Pe 0.68-0.69) for the increased dispersal scenarios under both the population-specific and generalised models. Although most values were close to zero, there was a Pe >0 for all individual populations under the population-specific model.

Under the generalised model, final metapopulation and individual population sizes for all scenarios were significantly different (Table 6.3). However, this was not the case for the population specific model, presumably due to a higher variance in input parameters.

Discussion

The predictions of the PVA models indicate the sensitivity of the metapopulation and local populations to changes in various parameters. In general, the existence of a metapopulation for H. albopunctatus would appear to increase the long-term persistence of this species, (compared to local populations) even with dispersal at low levels (1.67% in basic model). Modeling results showed that the long-term persistence of the generalized H. albopunctatus metapopulation was assured, with a Pe of 0, after 100 years. A number of scenarios, including a 55% reduction in offspring production and a decrease in the frequency of high rainfall years, did not impact the long-term persistence of the metapopulation. Similarly, the true model resulted in no metapopulation extinctions over 100 years. In contrast, the population specific model used individual input parameters for population size and carrying capacity based on mark-recapture and survey studies at each site. This model was still robust to changes in most parameters but responded sensitively to changes in juvenile survival.

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There are few comparable PVA models of amphibians with similar life histories.

Conroy and Brook (2003) used a stage-structured metapopulation model of two

Geocrinia species, although they did not specifically examine the impacts of variations in dispersal because this is extremely low in the species modeled. Griffiths and

Williams (2000) concluded that a metapopulation structure was essential to the long- term persistence of Triturus cristatus, with isolated populations that were unconnected by dispersal having a relatively high risk of extinction. In the current study, H. albopunctatus metapopulations seemed robust to catastrophic and demographic variations, even at low levels of dispersal and with low numbers of populations (e.g.

True model). The increased growth rate observed under the increased dispersal scenarios of all models, can be attributed to an influx of immigrants into each population, resulting in an accelerated growth rate for local populations and the metapopulation, compared to the base models. Although there was no risk of extinction for the metapopulation under the increased dispersal scenarios, there was a significantly increased probability of extinction for all local populations. This may be due to a high and variable turnover of populations due to an efflux of dispersers, akin to the patchy metapopulation model of Harrison (1991) or otherwise source-sink dynamics, with a few populations recruiting more successfully than an array of extinction-prone satellite populations (Harrison 1991).

Variations in juvenile survival had the greatest influence on metapopulation persistence under the population-specific model. For this and other models, field studies of H. albopunctatus (Chapter 5) revealed generally good adult survival although little information was obtained on juvenile survival despite an intensive mark-recapture study of 500 metamorphs. Conroy and Brooks (2003) also reported that the most sensitive life-history parameter in a PVA of Geocrinia spp., was juvenile survival. Besides the

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direct demographic importance of juveniles in becoming a reproductively active adult, juveniles may also be an important dispersal stage of the lifecycle. Adult anurans are often more vagile than juveniles. For example, Breden (1987) found that juvenile dispersal was responsible for most of the gene flow observed between populations of

Bufo woodhousei fowleri. Data on juvenile dispersal was not available for this PVA due to no recaptures of marked juveniles, however, adult dispersal was low (average of

1.67%). Genetic data (Chapter 2) indicated the presence of strong gene flow across the range of this species and even low levels of dispersal may be sufficient to eliminate genetic subdivision in this species. Given the relatively low dispersal capacity of adults, it may be inferred that tadpole or juvenile dispersal in times of flood may be important mechanisms driving gene flow and metapopulation structure. There is a need to quantify the importance of floods and tadpole and juvenile dispersal in this context.

Unfortunately, an accurate estimate of juvenile survival was not possible for H. albopunctatus, indicating a need for further attempts to quantify juvenile survival. The general literature estimate of juvenile survival used in the PVA may be an overestimate due to the more ephemeral and variable nature of H. albopunctatus ponds and the high rate of recruitment failure observed (Chapter 4). Poor recruitment will have ongoing implications for populations, as adults do not reach reproductive maturity until two years of age (Lee 1967). Consequently successive years of recruitment failure could result in potentially fatal outcomes for local populations. Regional metapopulations could also be affected if there was a high degree of environmental correlation between local populations and low rainfall was the primary cause of recruitment failure.

Fecundity was highly variable from year to year, in response to rainfall (Chapters 3,

4). Although changes in fecundity had an impact on population sizes and growth rates under all models, they did not adversely affect metapopulation persistence. Similarly, a

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decrease in the frequency of high rainfall years (in which reproduction and survival are favoured) to only one year in ten, did not affect metapopulation persistence. These data are consistent with several null models of amphibian behaviour (eg. Alford and

Richards 1999) and similar studies (Berven 1990; Pechmann et al. 1991) which have reported high levels of annual variation and multiple years of recruitment failure or reduced recruitment success, interspersed with occasional years of high rainfall and peak reproductive success (Conroy and Brook 2003).

In summary, the management implications of this PVA are that juvenile survival is critical to the long-term persistence of H. albopunctatus (see also Chapter 4). Further studies are required to gain an accurate picture of juvenile survival in the field. It is also apparent that the formation of a metapopulation comprising as few as five populations may be sufficient to buffer H. albopunctatus against extinction in the long-term. All metapopulations were robust to most changes in parameters and even with 95% adult mortality populations that were connected by dispersal were able to persist. Dispersal is particularly critical for amphibians in fragmented landscapes (e.g. Sjogren 1991) and the distance between populations is often a critical factor in maintaining regional metapopulations (Halley et al. 1996, Griffiths and Williams 2000). The life-history attributes of H. albopunctatus, including high fecundity, high adult longevity and low to moderate dispersal contribute to a robust regional metapopulation, responsive to changes, but with a strong chance of persistence over the long-term. To increase the accuracy and resolution of future analyses, further work is required to investigate the survival of juveniles due to the importance of this life-stage in regulating population dynamics.

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Table 6.1. Basic Model input parameters for Vortex simulations of H. albopunctatus population viability.

Parameter Estimate

Reproductive System Monogamous

First breeding age of males and females 2

Maximum Age 10

Maximum number of metamorphs per 37

year

Sex ratio at birth 1:1

Percentage males in breeding pool 100%

Average percentage of reproducing 100

females per year/site

Annual mortality as a % (SD)

Adult females 44.25 (38.79)

Adult males 38.32 (32.28)

Juveniles 77.5 (N/A)

Stable age distribution? No

Trend in K? No

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Table 6.2: Summary of input parameters for PVA.

Model Parameter Estimate(s)

Generalised Initial Population 87

Metapopulation Size

Dispersal 1.67%

Carrying Capacity 300

Number of 25

populations

Population-specific Initial Population 2 - 86

Size

Dispersal 1.67%

Carrying Capacity 4 - 300

Number of 25

populations

True Metapopulation Initial Population 12 - 86

Size

Dispersal 0-20%

Carrying Capacity 300

Number of 5

populations

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Table 6.3. Results of PVA simulation scenarios for H. albopunctatus. *=p<0.05 (compared with base scenario)

Model Scenario N Mean Mean population growth Probability of Mean Population Size (SD) determinis rate (SD) extinction tic r Local Meta Local Meta Local Meta Local Generalise 100 years 25 0.56 0.57 (0.36) 0.56 (0.39) 0 0 7439.54 (295.75) 297.58 (18.06) d Metapop. Worst case recruitment 25 0.52 0.48 (0.36) 0.48 (0.38) 0 0 7421.26 (328.34)* 296.85 (18.09)* Metamorph mortality at 25 0.51 0.47 (0.50) 0.46 (0.53) 0 0 7274.81 (766.41)* 291 (34.84)* 90% Metamorph mortality at 25 0.247 0.19 (0.57) 0.19 (0.63) 0.003 0.003 6572.91 263.71 (59.11)* 95% (1358.82)* Adult survival at 10% 25 0.737 0.68 (1.28) 0.68 (1.31) 0 0 5119.57 (999.04)* 204.78 (53.79)** 10 x dispersal 25 0.85 0.80 (0.44) 0.72 (0.55) 0 0.68 2115.4 (70.33)* 264.43 (14.44)* EV concordance 50% 25 0.56 0.58 (0.25) 0.56 (0.30) 0 0 7484.35 (196.81)* 299.37 (13.84)* High rainfall frequency 25 0.85 0.81 (0.43) 0.80 (0.45) 0 0 7477.64 (194.34)* 299.11 (15.09)* reduced (0.1) Population 100 years 25 0.85 0.74 (0.33) 1.39 (0.48) 0 0.062 1288.15 (90.19) 51.68 (6.99) Specific Worst case recruitment 25 0.52 0.425 (0.36) 1.03 (0.48) 0 0.063 1173.97 (165.26) 47.11 (9.8) Metamorph mortality at 25 0.25 -0.21 (1.046) 0.18 1 1 0 0 95% (0.941) Metamorph mortality at 25 0.51 0.34 (0.49) 0.76 (0.71) 0.077 0.139 858.23 (362.96) 37.6 (14.89) 90% Adult survival at 40% 25 0.81 0.70 (0.23) 1.35 0 0.067 1298.58 (53.44) 52.11 (5.93) (0.424) Adult survival at 10% 25 0.74 0.60 (0.82) 1.18 (0.86) 0 0.064 1077.14 (316.70) 43.24 (15.35) 10 x dispersal 25 0.85 0.78 (0.42) 1.17 (0.48) 0 0.69 897.76 (111.74) 112.27 (18.27) EV concordance 50% 25 0.85 0.76 (0.33) 1.43 (0.47) 0 0.061 1301.99 (59.46) 52.23 (6.15) High rainfall frequency 25 0.85 0.74 (0.34) 1.39 (0.49) 0 0.064 1289.36 (89.66) 51.73 (6.95) reduced (0.1) True 100 years 5 0.64 0.49 (0.37) 0.46 (0.41) 0 0.007 1453.11 (99.86) 292.71 (24.56)117

Chapter 7: Overview

The objective of this thesis was to investigate the current metapopulation structure of

Heleioporus albopunctatus, a frog species occurring in a now highly fragmented landscape that Main (1990) suggested had declined due to salinisation of its wheatbelt breeding sites. My studies revealed that H. albopunctatus populations were robust and had adapted successfully to the human-modified landscapes of the wheatbelt agricultural region. There were no detectable adverse impacts arising directly from habitat fragmentation although the choice of man-made breeding sites is of critical importance for the persistence of this species.

Examinations of genetic structuring (Chapter 2) revealed that there was gene flow across the range of this species, and no significant differences in genetic structure were detected between wheatbelt populations and those in continuous habitats. Although there was evidence of high levels of subdivision between some population pairs, this was not strongly related to either breeding site isolation or biogeographic barriers such as saline drainages. The choice of allozymes may have affected my ability to detect changes as allozymes may lack the sensitivity to detect contemporary patterns of gene flow in relation to habitat fragmentation. Without the use of more sophisticated techniques such as microsatellites (Love et al. 2004) it is difficult to differentiate observed patterns of structuring from historical patterns of gene flow.

Adverse impacts of habitat fragmentation were not detected in detailed studies of egg survivorship (Chapter 3). I found no detectable increase in mortality at smaller, more marginal breeding sites, and overall, H. albopunctatus egg clutches had very high survival to hatching.

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Studies of larval survivorship (Chapter 4) did not provide any evidence of the impacts of habitat fragmentation, but they did provide cause for concern due to widespread recruitment failure from pond drying. The fact that 86.11 % of breeding populations failed to recruit tadpoles during the study (Chapter 4) is indicative of the importance of local populations in driving metapopulation dynamics. A number of ponds produced metamorphs in each year of the study, in contrast to the majority of populations that failed to recruit, indicating a source-sink metapopulation model (see

Harrison 1991).

Further analyses of site quality indicated that hydroperiod was the most important factor driving metapopulation dynamics, with all ponds that failed to recruit either drying too early or failing to fill with water (Chapter 4). I also examined salinity of breeding ponds, and found that tadpoles were not present in pond with a conductivity measurement exceeding 5000 µS. However, the vast majority of pools were fresh and formed from rainwater and thus salinity is unlikely to have a large impact on tadpole development per se but rather, it may have a negative impact in a biogeographic sense by precluding breeding in low-lying (salinity-prone) areas of the landscape as discussed by Main (1990).

The ability of H. albopunctatus to switch to man-made breeding sites is a successful adaptive strategy allowing this species to persist in a highly modified landscape.

Ultimately, however, there are a number of repercussions arising from this strategy.

Breeding sites generally comprised salinity interceptor banks and roadside ditches, as well as damp areas of paddocks with no surface water. These sites may have been detected by cues such as soil moisture, but act as ecological traps since free-standing water is either never present or pond hydroperiod is too short. Obviously some breeding sites do have the right characteristics for successful recruitment and are thus regionally

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important for ensuring the persistence of this species at a landscape level. It is suggested that future research should focus on comparing “natural” (ie. continuously vegetated) breeding habitats with those in human-modified landscapes to provide a benchmark for assessing the impact of habitat modification and fragmentation on recruitment success.

Similarly, larval studies should be replicated at the broader landscape level to investigate the pond quality characteristics that determine recruitment.

Adult survival studies demonstrated good adult survival between years and a high site-fidelity (Chapter 5). Actual population sizes estimated using mark-recapture models were small and variable, but this appeared to relate to site quality rather than any direct impacts of habitat fragmentation. Mark-recapture studies revealed that dispersal between local populations occurred at low levels (1.67% on average), but this may be sufficient to ensure gene flow at a regional scale. Further mark-recapture studies would establish the nature and importance of dispersal between populations including the importance of flood-mediated dispersal of larvae in high rainfall events. Radio-tracking of adults would also be useful to establish non-breeding habitat use. If H. albopunctatus requires areas of native vegetation for shelter and foraging, the identification and protection of these areas may be equally important on a regional scale.

Finally, I consolidated all life-history data for this species and performed a population viability analysis (PVA) to determine the probability of extinction of local populations and metapopulations of this species in the central wheatbelt (Chapter 6).

PVA revealed that local populations and the metapopulation were very robust in the longer term (100 years), with all populations surviving extinction under the basic scenario. A number of other scenarios were run to investigate the impacts of demographic changes and catastrophes on populations. Although populations were robust to nearly all changes in parameters, the most realistic model (incorporating

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varying population sizes and carrying capacities) was sensitive to changes in juvenile survival (but not the number of offspring). Since estimates of juvenile survival were unavailable during mark-recapture studies (due to no recaptures of marked metamorphs) this parameter requires further study for more accurate estimation. Studies on other frog species have shown juveniles to be an important dispersal stage and responsible for most of the gene flow between populations (e.g. Breden 1987). If this is the case for H. albopunctatus, juvenile survival (and hence also recruitment success) is the most critical determinant of the species’ persistence. Further studies are needed to establish the survivorship and importance of juveniles to metapopulation dynamics.

In summary, I established that H. albopunctatus persists as a metapopulation with gene flow and adult dispersal between populations. Populations were robust with strong adult survival and adults displaying site fidelity. The species has successfully adapted to breeding in a human-modified landscape but the nature and hydroperiod of breeding sites is critical to their continued persistence. Management actions may be required to preserve successful breeding ponds and enhance breeding sites on a regional scale.

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Appendix 1: Publications relating to this study

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