Habitat Requirements for the Reintroduction and Persistence of the Northern Eastern Bristlebird (Dasyornis Brachypterus)

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Habitat requirements for the reintroduction and persistence of the Northern Eastern Bristlebird (Dasyornis brachypterus)

Zoë Leigh Stone MSc., BSc.

A thesis submitted for the degree of Doctor of Philosophy at The University of Queensland in 2018

School of Earth and Environmental Science

Abstract

The northern population of the eastern bristlebird (Dasyornis brachypterus, hereinafter referred to as the northern bristlebird) is critically endangered. It occupies grassy sclerophyll forest near the Queensland-New South Wales border. With an estimated wild population of only 38 birds and an 80% decline the population since monitoring began in 1980s, an intensive management strategy is currently underway which aims to bolster the population through reintroductions and maintain and restore suitable habitat for their persistence. Reintroductions are becoming a common technique for the conservation and management for a range of threatened species. Their chances of a successful release can be greatly increased with good ecological knowledge of the species concerned and the critical elements of habitat it requires, yet these are often lacking. Fundamental ecological information on habitat, food availability and threatening processes are important for understanding not only in the immediate survival of reintroduced individuals, but also the long-term persistence and recovery of populations. For the conservation and future reintroduction of northern bristlebirds to be successful, a good understanding of what habitat characteristics are necessary for persistence and how to maintain them is crucial.

To understand what habitat characteristics are associated with northern bristlebird persistence I conducted a detailed analysis of habitat structural and dietary resources, and the fire history, of northern bristlebird habitat across their historical range. To provide a background to this research I first conducted a review of the available published and unpublished literature to outline the current ecological knowledge about the northern bristlebird and establish hypotheses for field testing. Second, I investigated whether northern bristlebird presence was associated with particular vegetation structural characteristics within their grassy sclerophyll forest habitat. Third, I examined whether potential food limitation may be influencing northern bristlebird persistence, and whether preferred habitat elements were associated with specific invertebrate assemblages. Lastly, fire is a historically important disturbance within grassy forest habitat in my study region, and I tested whether fire regimes influence the extent and condition of northern bristlebird habitat.

I found little difference in plant community composition between current and historically occupied northern bristlebird sites. Northern bristlebird presence was, however, strongly influenced by the size of grassy sclerophyll forest habitat patches. Large habitat patches with an intact grassy understorey layer were more likely to contain northern bristlebirds. In addition I found that northern bristlebird presence was more probable at sites with dense cover of tall grass. The association of northern bristlebird presence with large habitat patches was highly contingent on the presence of these structural characteristics: large sites without high quality grassy understories were unlikely to have northern bristlebirds still present.

As a predominantly insectivorous species there has been some concern that food limitations within northern bristlebird habitat may be present. I found little evidence of differences in epigaeic (surface dwelling) invertebrates across sites, with similar abundance and biomass of invertebrates across current and historically occupied sites. There was also no significant variation in the composition of invertebrate communities between current and historically occupied sites. I found little variation in the abundance or composition of invertebrates between the seasons sampled, suggesting northern bristlebird habitat patches may provide fairly stable conditions for invertebrates. Invertebrate abundance and nutritional value corresponded with particular habitat characteristics. Places with tall thick grass structure and that were closer to the rainforest margin had more large invertebrates and were more nutritionally valuable. All northern bristlebird sites are within 400 m of rainforest habitat, and this may be at least partially because sites nearer to rainforest had more nutritious invertebrate resources.

Finally, I found that grassy sclerophyll forest patches for the northern bristlebird had substantially declined in extent over the last 30 years. Comparison of habitat extent between 1981 and 2009 showed a 50% decline in size of suitable grassy forest habitat. This decline was strongly associated with longer mean fire intervals and prolonged periods of fire absence. Longer fire intervals were not only strongly associated with a decrease in the extent of grassy habitat but also the condition of the grassy understorey. Grass volume and tussock cover declined with increasing fire intervals and prolonged fire absences. Given the strong relationship between northern bristlebird presence and size of habitat

patches, the substantial reduction in habitat patch size and grassiness over the last 30 years appears likely to be the main contributor to the northern bristlebird decline. Restoring appropriate fire regimes into northern bristlebird habitat is, therefore, a key management actions needed to maintain healthy habitat for bristlebird persistence.

The findings from my thesis provide critical insights into the habitat elements and fire regimes necessary for northern bristlebird persistence. I have shown that northern bristlebird persistence is highly dependent on large patches of grassy sclerophyll forest habitat, and that these patches are sensitive to modified fire regimes. A substantial decline in the extent of sclerophyll forest with thick, tall grassy understorey is likely the main contributor to their decline. Maintenance of this critical habitat by restoring more frequent fire to it will be crucial for the conservation and recovery of the northern bristlebird. This research has important implications for applied conservation management of these grassy forests and the recovery and future reintroduction attempts of the northern bristlebird. It also highlights the importance of conducting pre-release habitat assessments to determine proper management strategies for restoring and maintaining conditions for survival and long-term persistence of target species. Without understanding how and why a species has declined, halting such declines is unlikely, let alone sustaining the persistence of reintroduced individuals.

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Declaration by author

This thesis is composed of my original work, and contains no material previously published or written by another person except where due reference has been made in the text. I have clearly stated the contribution by others to jointly-authored works that I have included in my thesis.

I have clearly stated the contribution of others to my thesis as a whole including statistical assistance, survey design, data analysis, significant technical procedures, professional editorial advice, and any other original research work used or reported in my thesis. The content of my thesis is the result of work I have carried out since the commencement of my research higher degree candidature and does not include a substantial part of work that has been submitted to qualify for the award of any other degree or diploma in any university or other tertiary institution. I have clearly stated which parts of my thesis, if any, have been submitted to qualify for another award.

I acknowledge that an electronic copy of my thesis must be lodged with the University Library and, subject to the policy and procedures of The University of Queensland, the thesis be made available for research and study in accordance with the Copyright Act 1968 unless a period of embargo has been approved by the Dean of the Graduate School.

I acknowledge that copyright of all material contained in my thesis resides with the copyright holder(s) of that material. Where appropriate I have obtained copyright permission from the copyright holder to reproduce material in this thesis.

Publications during candidature

Accepted Papers:

 Chapter 3: Stone, Z. L., Tasker, E., & Maron, M. (2018). Grassy patch size and structure are important for northern Eastern Bristlebird persistence in a dynamic ecosystem. Emu - Austral Ornithology, 1-12

Submitted Papers:

 Chapter 4, submitted to Austral Ecology: Stone, Z. L., Tasker, E. & Maron, M. Patterns of invertebrate food availability and the persistence of an insectivore on the brink.

To be submitted:

 Chapter 5, to be submitted to International Journal of Wildland Fire: Stone, Z. L., Tasker, E. Maron, M. Frequent fires are needed to maintain grassy habitat of the northern eastern bristlebird.

Conference abstracts:

 Stone, Z. L., Tasker, E. and Maron, M. The role of fire disturbances on the distribution of structural resources for the Eastern Bristlebird. ICCB conference, Montpellier, August 2015  Stone, Z. L., Tasker, E. and Maron, M. Grass, bugs and fire: Habitat dynamics of the Northern Eastern Bristlebird. QOC conference, July 2016  Stone, Z. L., Tasker, E. and Maron, M. Burning an endangered species: importance of understanding habitat dynamics for Northern Eastern Bristlebird conservation. SEQ Fire and Biodiversity Consortium conference, September 2016  Stone, Z. L., Tasker, E. and Maron, M. Understanding habitat requirements for successful reintroduction and persistence of the Northern Eastern Bristlebird (Dasyornis brachypterus). AOC conference, November 2017  Stone, Z. L., Tasker, E. and Maron, M. One size doesn’t fit all: small populations provide important information for threatened species management. ECOTAS conference, November 2017

Contributing author publications:

None v

Publications included in this thesis

Chapter 3: Stone, Z. L., Tasker, E., & Maron, M. (2018). Grassy patch size and structure are important for northern Eastern Bristlebird persistence in a dynamic ecosystem. Emu - Austral Ornithology, 1-12.

Contributor Statement of contribution

Zoë Stone (Candidate) Conception and design (80%)

Analysis and interpretation (90%)

Drafting and production (80%)

Martine Maron Conception and design (10%) Analysis and interpretation (5%)

Drafting and production (10%)

Elizabeth Tasker Conception and design (10%) Analysis and interpretation (5%)

Drafting and production (10%)

Manuscripts included in this thesis

Chapter 4: Stone, Z. L., Tasker, E. & Maron, M. Patterns of invertebrate food availability and the persistence of an insectivore on the brink. Submitted to Austral Ecology

Contributor Statement of contribution

Zoë Stone (Candidate) Conception and design (80%)

Analysis and interpretation (90%)

Drafting and production (80%)

Martine Maron Conception and design (10%) Analysis and interpretation (5%)

Drafting and production (10%)

Elizabeth Tasker Conception and design (10%) Analysis and interpretation (5%)

Drafting and production (10%)

Chapter 5: Stone, Z. L., Maron, M & Tasker, E. Frequent fires are needed for maintaining grassy forest habitat for the northern bristlebird. To be submitted to International Journal of Wildland Fire.

Contributor Statement of contribution

Zoë Stone (Candidate) Conception and design (80%)

Analysis and interpretation (80%)

Drafting and production (80%)

Martine Maron Conception and design (10%) Analysis and interpretation (10%)

Drafting and production (10%)

Elizabeth Tasker Conception and design (10%) Analysis and interpretation (10%)

Drafting and production (10%)

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Contributions by others to the thesis

Two jointly published or submitted papers form part of this thesis (Chapters 3 & 4). The work for these articles was undertaken in majority by me, including generating ideas, data collection and analysis, interpretation and drafting of manuscripts. Martine Maron and Elizabeth Tasker contributed to generating ideas for research, provided statistical advice, and reviewed and commented on draft manuscripts.

Statement of parts of the thesis submitted to qualify for the award of another degree

None

Research Involving Human or Animal Subjects

No animal or human subjects were involved in this research

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Acknowledgements

Working in threatened species conservation, like finishing a thesis, can often seem daunting. With such large challenges for biodiversity, and many species clinging to existence the question of why, and what’s the point start to hang in the air. Sure, we can talk about the ecological value of these species and why biodiversity matters, which are valid points in themselves, but for conservation efforts to endure we need to inspire and evoke passion in others. So, to those who have inspired me, thank you. From an ever- charming presenter that narrated my youth, to familiar heroes and mentors at home who have fostered my intrigue in the rare and magical.

Thank you to people like Marie Taylor and Mick Clout, who have nurtured my interest and inspired me with their love of land and stories of restoration and conservation. Thank you to my fellow conservation friends and peers for taking up this challenge with me. Annette Evans and Esther Dale for their amazing friendship from undergrad to Masters, and now through our respective PhDs. To new friends and colleagues at UQ thank you for your support and general comradeship over the last four years. Particularly, thanks to Will Goulding for his friendship and endless conversations and through this PhD journey.

I would also like to thank all those academics that have supported and guided me through this process. Special thanks are needed for my supervisors Martine Maron and Liz Tasker for taking me up on this quest to figure out how we can save a rare, small brown bird I had never heard of before. I am grateful for all the guidance they have given me along the way, and for their supportive teaching when I needed a hand.

This project was supported and funded by the Eastern Bristlebird Recovery Northern Working Group, of whom I am extremely grateful for their support. From the team I would like to thank Lynn Baker, Liz Gould, Dave Charley, Sheena Gillman, Penny Watson, David Stewart, Steve King, Wil Buch, Allison Beutel and Anthony Molyneux for their dedication and continual support of myself and bristlebird conservation. Thank you to our passionate landowners, who have been incredibly helpful throughout this research. Extreme thanks go to Hugh and Elizabeth Starkey, Dave and Janet Murray, Leon Blank, and Ken and Jan ix

Drynan for being wonderful hosts, and for helping me up the hill when motor support was called for.

Traversing grassy forests and riffling through leaf litter in Australia can be challenging. I would like to thank my volunteers Jeff Hanson, Zeb Stone, Lesley Stone, Gwen Iacona, Chris O’Bryan and Michelle Gibson for not being too (visibly) disappointed at not seeing eastern bristlebirds and surviving the ticks and snakes with me. Thank you to Anita Cosgrove for her assistance in the lab sorting and identifying invertebrate samples.

Completing a PhD is hard, particularly when you are away from family. People often ask if you can see light at the end of the tunnel yet. The PhD journey can seem like a long tunnel that bends and meanders through unexpected terrain which seems to continue endlessly in darkness. Yet that tunnel will always end. At some point, there is a faint glow on the horizon, that brightens with every field plot surveyed, extra paragraph written, or analysis completed. The light might creep up unexpectedly, around one last bend that obscured it to seem like it was never there. Thank you so much to my partner Jeff Hanson, who navigated the tunnel with me and held my hand through it all.

Finally, I am profoundly grateful to my family. They have inspired me through their love and passion of doing and creating with a purpose. That taught me to hope and to not accept the frivolous, to look beyond the cynicism and selfishness. Thank you to my Mother, Lesley, for introducing me to science and helping see the wonder in asking questions, finding the answers or at least creating a new question. Thank you for being my constant cheer leader. Thank you to my brother Zeb, for being one of the best people I know, tirelessly trying to build a better society that cares about more than the things they own. Thanks for taking care of the people side of things while I look to the land.

Lastly to my Dad, my Alexander the great. Thank you for telling stories. For spouting poetry from tabletops with joy and passion. You have taught me how to love, how to laugh and how to revel in the magic. Thank you for showing me how to capture the world on paper, whether through writing, photography or drawing. Tunnels can be dark places, but you taught me to look around for the spiders and glow-worms, listen for running water and

catch droplets from mossy walls. I hope your tunnel isn’t too long and you find your way out soon. But until then look up for skylights, climb over the boulders and follow the wind. Here is a story I wrote, it’s about frisky birds and fire. A story of hope and resilience in a grassy forest of Australia. This story is for you.

‘And we danced, on the brink of an unknown future, to an echo of a vanished past’

John Wyndham, Day of the Triffids

Financial support

This research was supported by an Australian Government Research Training Program Scholarship, the New South Wales Environment Trust, the Eastern Bristlebird Recovery Northern Working Group, the NSW Office of Environment and Heritage, the Saving our Species program of the NSW Office of Environment and Heritage; Birds Queensland; Birdlife Australia; and The University of Queensland.

Keywords eastern bristlebird, threatened species, conservation, reintroduction, habitat selection, invertebrates, fire regimes, management, sclerophyll forest

Australian and New Zealand Standard Research Classifications (ANZSRC)

ANZSRC code: 050211, Wildlife and Habitat Management

ANZSRC code: 050104, Landscape Ecology

ANZSRC code: 050202, Conservation and Biodiversity

Fields of Research (FoR) Classification

FoR code: 0501 Ecological Applications

FoR code: 0502 Environmental Science and Management

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Table of Contents

Abstract ...... i

Declaration by author ...... iv

Publications during candidature ...... v

Publications included in this thesis ...... vi

Manuscripts included in this thesis ...... vi

Contributions by others to the thesis ...... viii

Statement of parts of the thesis submitted to qualify for the award of another degree ..... viii

Research Involving Human or Animal Subjects ...... viii

Acknowledgements ...... ix

Financial support ...... xii

Keywords ...... xii

Australian and New Zealand Standard Research Classifications (ANZSRC) ...... xii

Fields of Research (FoR) Classification ...... xii

List of Figures ...... xvii

List of Tables ...... xx

List of Plates ...... xxi

Appendix ...... xxi

Chapter 1 - Introduction ...... 1

1.1. Background to the problem ...... 1 1.1.1. Reintroduction biology ...... 2 1.1.2. Reintroduction of the northern eastern bristlebird ...... 4 1.1.3. Factors influencing habitat suitability for species persistence and reintroduction attempts ...... 5 1.1.4. Habitat requirements of the eastern bristlebird ...... 15 1.2. The problem...... 16 1.3. Thesis aims & objectives ...... 17

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1.4. Thesis structure ...... 19 Chapter 2 - Review of the ecology of the northern population of the eastern bristlebird ...... 23

2.1. Introduction ...... 24 2.2. Methods ...... 26 2.2.1. Literature review ...... 26 2.2.2. Northern bristlebird distribution and population trends ...... 27 2.2.3. Habitat characteristics ...... 29 2.3. The eastern bristlebird ...... 30 2.3.1. Southern eastern bristlebird ...... 30 2.3.2. Northern eastern bristlebird ...... 32 1.3.3. Potential habitat dynamics of the northern bristlebird ...... 42 Chapter 3 - The importance of grassy patch size and structure for northern Eastern Bristlebird persistence in a dynamic ecosystem ...... 47

3.1. Abstract ...... 48 3.2. Introduction ...... 48 3.3. Methods ...... 52 3.3.1. Study Area ...... 52 3.3.2. Habitat assessments ...... 53 3.3.3. Presence/absence data ...... 54 3.3.4. Spatial characteristics ...... 54 3.3.5. Statistical analysis ...... 56 3.4. Results ...... 59 3.4.1. Plant community composition ...... 59 3.4.1. Habitat attribute models ...... 59 3.4.2. Classification tree ...... 61 3.5. Discussion ...... 62 3.6. Acknowledgements ...... 66 Chapter 4 - Patterns of invertebrate food availability and the persistence of an insectivore on the brink ...... 67

4.1. Abstract ...... 68

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4.2. Introduction ...... 68 4.3. Methods ...... 71 3.3.1. Study area ...... 71 3.3.1. Northern bristlebird presence and absence...... 71 3.3.2. Invertebrates ...... 72 3.3.3. Habitat patch characteristics ...... 75 3.3.4. Statistical analysis ...... 75 4.4. Results ...... 77 3.4.1. Bristlebird presence in relation to invertebrate resources ...... 77 3.4.2. Invertebrate patterns within bristlebird habitat patches ...... 79 4.5. Discussion ...... 83 4.6. Acknowledgements ...... 87 Chapter 5 - Frequent fires are needed for maintaining grassy forest habitat for the northern bristlebirds...... 88

5.1. Abstract ...... 89 5.2. Introduction ...... 89 5.3. Methods ...... 91 5.3.1. Study area ...... 91 5.3.2. Northern bristlebird presence/absence ...... 93 5.3.3. Fire history ...... 93 5.3.4. Vegetation characteristics ...... 94 5.3.5. Invertebrate resources ...... 94 5.3.6. Habitat patch extent and canopy cover ...... 95 5.3.7. Statistical analysis ...... 97 5.4. Results ...... 98 5.4.1. Change in habitat extent between 1981 and 2009 ...... 98 5.4.2. Role of fire history and change in habitat extent on northern bristlebird presence ...... 99 5.4.3. Influence of fire on habitat extent and condition ...... 100 5.5. Discussion ...... 102 5.5.1. Contraction of grassy patches ...... 103 5.5.2. Altered habitat condition ...... 104 xv

5.5.3. Conservation implications ...... 105 Chapter 6 - Thesis conclusion & reccommendations ...... 107

6.1. Thesis overview ...... 108 6.1. Contributions of this thesis to scientific knowledge ...... 111 6.1.1. Fire ecology and vegetation dynamics ...... 112 6.1.2. Adaptive management of grassy forests ...... 1122 6.1.3. Invertebrate ecology ...... 113 6.1.4. Reintroduction biology ...... 116 6.1.5. Northern bristlebird ecology ...... 117 6.2. Northern bristlebird management recommendations ...... 118 6.2.1. Habitat management ...... 118 6.2.2. Reintroductions ...... 123 6.3. Study limitations ...... 124 6.4. Future research priorities ...... 125 6.5. Conclusion ...... 126 Appendix ...... 128

References ...... 133

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List of Figures

Figure 1.1. The three main components that contribute to reintroduction success in threatened species conservation. Examples of the factors that affect initial establishment of individuals and the long-term persistence of reintroduced populations are given, along with supporting references for each...... 7

Figure 1.2. An overview of the thesis structure...... 18

Figure 2.1. Historical and recent records of eastern bristlebird across south-east Australia. Adapted from Baker (1997), Lamb (1993) and Clarke & Bramwell (1998)...... 24

Figure 2.2. Conceptual model of the habitat dynamics occurring within heath habitats of the southern and central eastern bristlebird populations in Victoria and New South Wales...... 31

Figure 2.3. Population trends fitted using log-linear models for northern bristlebird since 1982 based on number of individual birds recorded during official surveys (squares, dotted trend line) and population estimates (circles, solid trend line) which includes additional observations from unofficial surveys. Based on monitoring reports and unpublished data between 1982 and 2016 (data from reports listed in Appendix 2)...... 32

Figure 2.4. Example of various states of northern bristlebird habitat: Column 1) thick grassy sites that are currently occupied; column 2) occupied sites that have been recently burnt by fire (top) or overgrazed (bottom); and column 3) abandoned sites with increased cover of rainforest and shrub species...... 36

Figure 2.5. Example of grassy forest transitioning into shrubby rainforest at a historically occupied northern bristlebird site in the Border Ranges (Site BL1, Richmond Gap). Photo points were taken at the same location facing north (top), south and west (bottom) in 1985, 1993 and 2013. Photos supplied by Dave Charley (Wildsearch Environmental Services). 39

Figure 2.6. Conceptual model of potential habitat dynamics occurring within northern bristlebird habitat based on expert knowledge and monitoring data...... 43

Figure 3.1. Historic and current distribution of the eastern bristlebird along the south-east coast of Australia with insert showing all known historic records for the northern population in QLD and northern NSW. Adapted from Baker (1997)...... 53 xvii

Figure 3.2. Two-dimensional ordination plot of plant community composition of currently (black) and historically occupied (grey) northern eastern bristlebird sites classified by region...... 60

Figure 3.3. Classification tree showing habitat attributes that explain northern eastern bristlebird presence. The tree is read by following the left split if statement is true and following the right split if statement is false. Each node box shows the probability of northern bristlebird presence. For example, 40% of our data (15 sites) had <0.31% Crofton weed cover (1st left split), and a 16% probability of northern bristlebird being present...... 62

Figure 4.1. Clockwise from top left: (a) historic range of the eastern bristlebird; (b) all known territories (recorded as centroid points) of the northern population of the eastern bristlebird in Queensland and northern New South Wales; (c) example of plot layout for two sites within a single grassy habitat patch where vegetation and invertebrate sampling were conducted, the former along two transects at right-angles to each other, the latter within randomly sampled plot subdivisions (pitfall traps = circles, leaf litter samples = crosses)...... 72

Figure 4.2. Two-dimensional ordination plots of (a) overall, (b) autumn and (c) spring invertebrate community composition for current (black) and historically occupied (grey) northern bristlebird sites classified by region...... 79

Figure 4.3. Differences in invertebrate abundance, biomass, and nutritional value between autumn and spring sampling seasons for all current (black) and historically occupied (grey) northern bristlebird sites...... 80

Figure 4.4. Relationship between (a) distance to rainforest edge and (b) habitat patch size on macro-nutritional value (mg of crude protein + fat) based on five main invertebrate groups present within northern bristlebird sites…………………………………….……….…80

Figure 4.5. Total abundance of (a) macro and (b) micro invertebrates captured in northern bristlebird sites in relation to grass height…………………………………………………...…80

Figure 5.1. (a) Historical and current records of the northern bristlebird, showing (1) Conondale, (2) Main Range, (3) Lamington and (4) Border Ranges regions in which changes to grassy sclerophyll forest habitat extent between (b) 1981 (c) and 2009 were mapped (detail shown for a Conondale example) ...... 92

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Figure 5.2. Example of habitat characteristics used to delineate northern bristlebird habitat: centre) differences in canopy structure between vegetation types, a & b) examples of clumping native tussocks, c) light green of weed infested understorey and d) sclerophyll forest with shrubby mid-stratum...... 96

Figure 5.3. a) Average size of grassy sclerophyll forest habitat patches and (b) percent of canopy cover within patches across all northern bristlebird sites between 1981 and 2009.*** denotes p < 0.001, while NS. represents no significance (p > 0.05) according to t- test between years...... 99

Figure 5.4. Change in habitat extent and canopy cover between 1981 and 2009 for current (‘1’) and historically occupied (‘0’) northern bristlebird sites. Dotted line represents zero change in habitat patches between time frames while significance based on paired t-test between 1981 and 2009 is displayed above...... 100

Figure 5.5. The relationship between time since fire on (a) grass volume (height x cover), and (b) cover of the dominant grass species (%) in all bristlebird sites...... 102

Figure 6.1. Updated habitat dynamics model (from Chapter 2) for the northern bristlebird based on findings from Chapters 3-5 of this thesis. Stars denote information that has been added based on chapter results...... 122

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List of Tables

Table 2.1. Threatened species listed under the NSW Biodiversity Conservation Act (2016) recorded at northern eastern bristlebird locations. Adapted from Sandpiper Ecological Surveys (2000)...... 35

Table 2.2. Common plant species found within northern bristlebird territories. Figures are percent cover in territories originally identified by Holmes (1989). More recent figures are from Lamb (1993) and Stone (2013). 1993 and 2013 surveys focussed on changes in understorey vegetation therefore no tree data is available...... 38

Table 3.1. Description of northern bristlebird grassy habitat and non-habitat used to define habitat patches. Adapted from Rennison (2016)...... 55

Table 3.2. Habitat patch variables used for predicting northern eastern bristlebird presence across their historic range...... 58

Table 3.3. Relative performance of six models, representing five alternative hypotheses about the drivers of northern eastern bristlebird persistence across their historical range. 61

Table 4.1. Number of invertebrates collected in pitfall traps and leaf litter samples within northern bristlebird habitat patches...... 78

Table 4.2. Relative performance of alternative models testing northern bristlebird presence in response to invertebrate predictors in currently occupied and historically occupied sites...... 79

Table 4.3. Relative performance of alternative models testing invertebrate abundance, biomass and nutritional value response to habitat patch characteristics across currently occupied and historically occupied northern bristlebird sites...………………………………79

Table 5.1. Relative performance of alternative models for northern bristlebird persistence in response to fire history and change in habitat extent between 1981 and 2009...... 99

Table 5.2. Relative performance of alternative models testing influence of fire history (1981-2009) on grass volume, tussock cover, dominant grass cover, invertebrate biomass, and change in habitat extent within all bristlebird sites. Model coefficients are time since fire (TSF), mean fire interval (MFI) and longest inter-fire interval (LFI)...... 101

List of Plates

Plate 1. Fire edge between recently burnt and unburnt grassy habitat in the Border Ranges...... 23

Plate 2. Thick grassy sclerophyll forest habitat of an occupied northern bristlebird site in the Border Ranges...... 47

Plate 3. Invertebrate pitfall trap beneath dense Sorghum tussock at Spicers Gap, Main Range...... 67

Plate 4. Planned burn at an occupied northern bristlebird site in the Border Ranges...... 88

Plate 5. Captive northern bristlebird singing at Currumbin Wildlife Sanctuary...... 107

Appendix

Appendix 1. List of known northern bristlebird locations with number of territories defined by Holmes (1989) the number of sites (grouped territories) present, and the number of sites surveyed for this thesis...... 128

Appendix 2. The number of individual birds recorded and population estimates during systematic surveys between 1982 and 2016 for the northern bristlebird across known and new sites. Population trend (Fig 3.) is based on the number of birds recorded (*) and reported in official monitoring reports, despite higher population estimates available because of additional or spontaneous surveys from personal records of Eastern Bristlebird Recovery Northern Working Group members that have been added to official counts. ... 129

Appendix 3. Correlogram showing spatial autocorrelation observed in the model residuals for the null hypothesis when the spatial autocovariate variable is not included (grey) compared to null hypothesis with spatial autocovariate variable included (black)...... 130

Appendix 4. Fire history information used to model the influence of fire on northern bristlebird habitat characteristics at current and historically occupied sites...... 131

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Chapter 1

Introduction

1.1. Background to the problem

Endangered species recovery has created some of the most inspirational conservation stories around the globe. Iconic, charismatic species are often associated with high-profile conservation efforts and can lead to important outcomes: the establishment of large protected areas (Walpole & Leader-Williams, 2002); eradications of invasive species from whole islands or landscapes (Towns & Broome, 2003), or long-term, intensive programmes that recover the most critically endangered of species (Jachowski & Lockhart, 2009; Jones et al., 1995). Although single-species conservation has been criticised (Lindenmayer et al., 2007; Roberge & Angelstam, 2004; Simberloff, 1998), it can not only conserve the targeted species, but also confer benefits to non-target species and contribute to overall ecosystem health (Baber et al., 2009). For instance, many conservation programmes that target introduced predators to benefit particular prey species, exemplified by eradication of mammalian predators on off-shore islands, confer significant benefits to a range of additional native and endemic species (Jones et al., 2016).

Around the world, various levels of success have been achieved for the recovery of endangered species. In some cases, species have returned from populations consisting of only a few breeding individuals to become stable and viable populations. The Chatham Island Black Robin (Petroica traversi) for instance, which was reduced to a single breeding pair, now has a relatively stable wild population of 250 individuals as a result of intensive captive breeding and reintroductions on offshore island sanctuaries (Butler & Merton, 1992). In Mauritius, the endemic kestrel (Falco punctatus) was reduced to only four known pairs in the 1970s, but has since recovered to an estimated population size of 800 individuals thanks to an intensive recovery effort (Jones, 2008). In Australia, successful reintroductions of the noisy scrub bird (Atrichornis clamosus) in south-western Australia resulted in the founding population of 45 birds increasing to 770 individuals by 2001 (Burbidge et al., 2005).

In south-east Australia, the eastern bristlebird (Dasyornis brachypterus) has undergone substantial declines in extent and population size due to habitat loss from urbanisation, land clearance and changes to natural fire regimes (Baker, 1997; OEH, 2012). While the southern population has become stable thanks to habitat protection, threat management and successful translocations (Baker et al., 2012; OEH, 2012), the northern population hangs by a thread. Without immediate action this ecologically unique population will very likely be lost. Reintroduction of captive bred individuals to help reinforce the tiny population of only 38 estimated individuals will be crucial for their persistence. Before these reintroductions take place, important ecological information is needed for us to be confident in its success, particularly in terms of habitat requirements and fire management.

While conservation efforts like reintroductions can result in the successful recovery of highly endangered species, the potential for success will be enhanced if reintroductions are informed and accompanied by comprehensive research, monitoring and appropriate habitat management. This thesis aims to aid future habitat management and reintroduction by investigating the role of habitat requirements in northern bristlebird persistence. In this chapter I provide some background to the science of reintroduction biology and the necessary elements that promote persistence after release. I then relate this to the northern population of the eastern bristlebird and outline the objectives and structure of the chapters that follow.

1.1.1. Reintroduction biology

The field of reintroduction biology aims to improve understanding of the ecological processes that influence successful translocations or reintroductions (Armstrong et al., 2015; Sarrazin & Barbault, 1996). Translocation is broadly defined as ‘the intentional movement of organisms from one place to another to conserve species and restore ecosystems’ (Armstrong et al., 2015, page 1), and includes the reintroduction of individuals to previously occupied habitats (Armstrong et al., 2015; Armstrong & Seddon, 2008). Reintroduction biology has emerged from the long history of threatened species translocations, particularly in New Zealand and Australia where translocation attempts

began as early as 1895 (Seddon et al., 2015). Since then, translocations have become an important conservation tool for many threatened species, though many still fail (Fischer & Lindenmayer, 2000; Griffith et al., 1989; Seddon et al., 2007).

Reintroduction failure appears to be more likely when conservation projects fall short in their understanding of key threats and ecological requirements of the target species, as well as other interacting species within the receiving environment (Armstrong & Seddon, 2008; Griffith et al., 1989; Seddon, 2010). In a review of reintroduction attempts Fischer and Lindenmayer (2000) found that only 30% of the reintroductions described in the 116 papers they reviewed were classified as successful. They also found that most studies of unsuccessful reintroductions provided little or no mention of the cause of the target species' continued decline. When studies did provide causes, they tended to be those not anticipated or addressed prior to reintroduction (Fischer & Lindenmayer, 2000).

Reports of failed reintroductions are important because they demonstrate that intensive pre-reintroduction assessment of habitat suitability, potential ecological relationships and stochastic events are vital to the success of reintroductions, particularly in ensuring that the potential for long-term persistence is maximised. For example, Moorhouse et al. (2009) found that failed reintroductions of water voles (Arvicola terrestris) to the upper Thames Region in the UK were strongly associated with vegetation condition at release sites. Increased mortality of individuals was linked to lower vegetation cover, or quality of release site habitat. A failed reintroduction attempt in Australia of the brown treecreeper (Climacteris picumnus) to a nature reserve found that aggressive, non-target species (e.g. noisy miners, Manorina melanocephala) had a stronger impact on short-term establishment than did habitat condition itself (Bennett et al., 2012).

Reintroduction research often focuses on the establishment of the released individuals immediately following their release (Armstrong & Seddon, 2008; Taylor et al., 2017). A recent review by Taylor et al. (2017) found that only 32% of published papers that reported reintroduction attempts addressed population persistence at release sites with most studies only addressing population establishment, even though long-term persistence is

the overall objective of reintroduction actions (Seddon, 1999). Even if reintroductions are deemed ‘successful’ in the short term, the small, usually isolated and naïve nature of these populations means they are still highly susceptible to unforeseen disturbances and threats that may occur infrequently, or over longer time scales. For example, after the reintroduction of noisy scrub birds to SW Australia was deemed ‘successful’, uncontrolled wildfires in 2000 and 2005 reduced the recovering population from roughly 800 to 343 individuals (Burbidge et al., 2005; Comer et al., 2010). Without long-term monitoring these declines would not have been observed, and absence of long-term management may have meant failure in creating a self-sustaining population. In contrast, many plant species require periodic disturbance to maintain reintroduced populations. Disturbance can open space, alter competitive dynamics, and redistribute resources (Maschinski et al., 2012). Many plant species rely on disturbance events such as fire for regeneration (Monks et al., 2012), therefore the disturbance regime at a site can have a strong influence on the establishment and persistence of reintroduced populations (Maschinski et al., 2012). To ensure long-term success of reintroductions, good understanding of species’ requirements for both immediate establishment and long-term persistence are vital, as well as having adequate long-term monitoring and management that can detect issues and potentially respond to them (Armstrong & Seddon, 2008).

1.1.2. Reintroduction of the northern eastern bristlebird

Australia has undergone significant changes in land use and fire regimes since European arrival. These changes have often been catastrophic to native ecosystems and their wildlife, with many species now requiring intensive management actions to avoid extinction. One such species is the eastern bristlebird. The eastern bristlebird is currently listed as ‘Endangered’ on the IUCN Red List of threatened species due to its small range and severely isolated populations (Red List criteria B1ab + 2ab) (IUCN, 2017). The total estimated population size is only 2500 individuals, with around 90 % of individuals found within the stable central population (Baker, 1998b; OEH, 2012; Roberts et al., 2011). The southern population consists of roughly 380 individuals (Bramwell et al., 1992). The southern and central populations, hereafter grouped as the ‘southern bristlebird’, are part of the same management unit because of similar habitat types (Baker, 2000) a lack of

genetic distinction (Roberts et al., 2011) and similar threats/management needs (OEH, 2012).

However, the northern population is perilously close to extinction. Since monitoring began in the 1980s it has declined by over 80% (OEH, 2012; Wildsearch Environmental Services, 2016b). The 2016 annual census estimated a total wild population size of only 38 individuals, and with only five known breeding pairs the effective population size is minute (Wildsearch Environmental Services, 2016b). This decline of the already critically endangered northern population (hereinafter the ‘northern bristlebird’) has prompted intensive action by the Eastern Bristlebird Northern Recovery Working Group (henceforth the ‘Northern Working Group’). As part of the group's work, a captive breeding population was established in 2013 at Currumbin Wildlife Sanctuary with the aim to release individuals back into suitable habitat patches to help boost the wild population. The captive breeding has been highly successful and will soon be at a stage for reintroductions to the wild to take place (L. Baker pers. comm.). There is, however, little scientific understanding of what has driven the disappearance of northern bristlebirds from their remaining habitat, and the role fire plays in habitat maintenance. To address these issues, this thesis aims to reveal ecological information needed to support the successful reintroduction of northern bristlebirds. To establish the habitat and ecological requirements of northern bristlebird for long-term persistence, we need to understand what factors influence habitat suitability, and how conservation programmes can manage habitat for the benefit of reintroduced species.

1.1.3. Factors influencing habitat suitability for species persistence and reintroduction attempts

Improving our knowledge of key habitat requirements and species ecology is a vital part of reducing risks associated with reintroductions (Hoegh-Guldberg et al., 2008). We cannot expect intensive management actions such as captive breeding and subsequent reintroductions to be successful if release sites do not provide habitat that is capable of supporting species persistence (Armstrong & Seddon, 2008; Griffith et al., 1989). How wildlife interact or select habitat depends on spatial and temporal scales (Block & Brennan, 1993; Hall et al., 1997; Hutto, 1985). In reintroduction biology, habitat interactions often occur at the home range level, as human involvement (and anthropogenic modification of

landscapes) in the reintroduction process has largely predetermined habitat selection at the geographic scale. This is also particularly relevant for dispersal limited species who, following reintroduction will be restricted to pre-selection geographic ranges, and therefore the suitability of home range and site level habitat characteristics is an important component of reintroduction success. In ecology, habitat terms are often ambiguous, with various terms used inconsistently, and it is important for ecologists to define clearly these terms (Hall et al., 1997). For this thesis, habitat is defined as the physical environment in which species occupy, focussing on the vegetative structure, microclimatic conditions, soil, topography and resources (both dietary and structural). I have included non-threatening -in terms of long-term persistence of species - ecological interactions within this definition, and excluded threatening ecological interactions that have been altered by anthropogenic impacts (e.g. introduced predators or altered abundances from landscape modification). I define suitable habitat as habitat within a species geographic range that provides appropriate home range, site and microsite conditions to support a viable, breeding population. Variation in potential habitat (i.e. habitat at the geographic scale) between and within home ranges, and at the finer site scale of habitat all influences the suitability of a habitat patch for long-term persistence of natural and reintroduction populations.

The suitability of an area of habitat for reintroduction is determined by population, ecosystem (e.g., physical environment) and threatening processes that influence both the immediate establishment and long-term persistence of a species to different degrees (Armstrong, 2008). Population level factors are often of immediate importance for reintroductions, influencing establishment success and necessary release strategies, but can also have long-term impact on persistence (i.e. genetic makeup). Environmental and threat related inputs generally have ongoing effects on reintroduction success, and play a crucial part in determining persistence in the landscape. Of these environment level factors, the availability of 1) appropriate resources such as food or nest sites, 2) protection or refuge from threats such as predators, habitat loss or disease, and 3) landscape configuration of habitat to allow for self-sustaining populations, are key determinants of reintroduction success (Fig. 1.1).

Elements for successful reintroduction

POPULATION HABITAT THREATS

ESTABLISHMENT PERSISTENCE ESTABLISHMENT PERSISTENCE ESTABLISHMENT PERSISTENCE • Population size • Population size • Refuge availability • Breeding habitat • Predator abundance • Predator/parasite • Demography (wild & released) • Food abundance • Habitat structure • Disturbance prevalence • Individual condition • Connectivity • Local microclimatic or & configuration • Disease, pathogen • Stochastic events & behaviour • Genetic diversity seasonal conditions • Resource stability or parasite presence • Disturbance regimes • Release strategy • Ecological interactions • Habitat size • Environmental fluctuation Frankham et al. (2014) Albrecht & Maschinski (2012) Moseby & O’Donnell (2003) Steffens et al. (2005) Moseby et al. (2011) Urbanek et al. (2010) McPhee (2004) Miller et al. (2009) Moorhouse et al. (2009) Michel et al. (2010) Short et al. (1992) Comer et al. (2010) Gusset et al. (2006) Taylor et al. (2005) Scheepers & Venzke (1995) Howells & Edward-Jones Grey-Ross et al. (2009) Hamilton et al. (2010) Nelson et al. (2002) VonHoldt et al. (2008) Bennett et al. (2012) (1997) Brannelly et al. (2016) Holl & Hayes (2006)

Availability of structural and dietary resources for persistence

The most immediate requirement for target species following reintroduction is the availability of critical resources that are needed for survival. Critical resources include dietary resources or shelter that protect individuals from adverse conditions or predation (refuges). In modified habitats that do not support appropriate resource levels for survival, provision of supplementary food resources or addition of shelter refuges during reintroduction attempts can help alleviate limitations and assist in the establishment and increase reproductive success of individuals (Castro et al., 2003; Chauvenet et al., 2012; Schoech et al., 2008). Vegetation structure of habitat provides valuable resources for persistence. Species like the northern bettong (Bettongia tropica) rely on a thick grassy understorey for protection as well as the abundance of truffle food resources (Bateman et al., 2012). Reintroduction success of this species is therefore dependent on whether habitat provides appropriate vegetation and food resources. How and when species use such resources, and how they are distributed within habitat can strongly influence survival and reproduction rates. In the northern bettong example, truffle food resources fluctuate

with rainfall, and during drier periods cockatoo grass (Alloteropsis semialata) becomes a more important resource in the absence of truffles (Abell et al., 2006). For reintroductions, it is therefore important to release individuals into habitat that provides adequate truffle and cockatoo grass foraging opportunities, particularly with the increasing impact of climate change on drought and rainfall patterns (Abell et al., 2006).

Different habitat resource types may have varying levels of influence on habitat suitability and the fitness of a species (Chalfoun & Martin, 2007). For instance, habitat that has undergone moderate levels of overall degradation may be better at supporting species persistence than higher quality habitat that is missing a single critical resource. This can be exemplified by the loss of specific breeding sites or important refuges within what would otherwise be considered to be suitable habitat. For instance, reintroduction attempts of the western swamp tortoise (Pseudemydura umbrina) in Western Australia have involved pre-release habitat management to provide artificial aestivation tunnels, an activity that has proved crucial to the survival of released individuals through providing important refuge during against wildfires within marginal remnant habitat (Burbidge et al., 2010).

The most important factors driving habitat selection can differ depending on the spatial scale. Chalfoun and Martin (2007) demonstrated that Brewer’s sparrows (Spizella breweri) selected habitat differently depending on which spatial scale was assessed (landscape versus territory scales). Their study found food availability was more important at the broader landscape scale for selecting breeding habitat, while at the fine territory scale the density of shrubs for nesting played a more important role in selecting breeding sites (Chalfoun & Martin, 2007). This demonstrates the complex resource requirements of species within their habitat, with different resources having varying degrees of importance at different spatial scales.

Habitat suitability is not constant, with changes over time and space depending on local climatic, edaphic or disturbance conditions (Orians & Wittenberger, 1991). In breeding birds, spatial and temporal variability in food resources has been well shown to strongly influence reproduction (Martin, 1987, 1995). Reproductive consequences of food variability

can include changes to clutch sizes (Ricklefs, 1980; Zanette et al., 2006), offspring size and body condition (Heezik & Davis, 1990), and nesting attempts (Nagy & Holmes, 2005; Zanette et al., 2000). Home ranges can also spatially shift following temporal changes to resource distribution (Lurz et al., 2000; Schradin & Pillay, 2006). Temporal changes in resource distribution may be particularly problematic during reintroduction attempts if individuals need to move outside managed or protected habitat areas to acquire necessary resources.

For long-term persistence of species, habitat patches need to be able to support species in the presence of ever changing environmental conditions in which individuals and populations may have changing resource needs (George & Zack, 2001). In the short term, individuals have different resource requirements during different life stages across their lifespan and habitat needs to provide for each of these stages for successful survival. On top of this, environmental variation such as droughts lead to periods of resource stress or bottlenecks that reduce availability of critical resources for survival (Maron et al., 2015). For long-term persistence of populations, habitat resilience is crucial in allowing a range of suitable resources to support both individual needs and long-term population growth. Climate variation, e.g., drought, can hugely impact the availability of resources within ecosystems, and subsequently alter community assemblages (Smith, 1982). Drought conditions in Borneo in 1997-1998 led to substantial reductions in fruit availability for sun bears (Helarctos malayanus) (Fredriksson et al., 2007), and were found to cause the local extinction of a fig pollinating wasp because of reduced inflorescence production of its host fig species (Harrison, 2000). Such drastic changes in resource availability can have strong cascading effects on species that rely on the resources in question. For insectivorous species (e.g., the eastern bristlebird) variation in invertebrate availability as a result of temporal changes in climate can strongly impact on their ability to persist (Williams & Middleton, 2008). Therefore reintroductions of such species need to occur within habitat that is resilient or heterogeneous and capable of providing suitable alternative prey during periods of environmental stress.

Protection from threats

Another key attribute that determines the suitability of habitat for species reintroduction is the presence (or absence) of threatening processes or the original causes of the species’

disappearance. Globally, recent species declines have largely been in response to anthropogenic impacts that have caused major habitat loss or modification, direct exploitation, introduction (or change in distribution/behaviours) of novel predators, diseases and pathogens (Loehle & Eschenbach, 2012). Reintroduction attempts are likely to fail if release sites are still subject to these threats (Loehle & Eschenbach, 2012).

i. Habitat modification - Disturbances and stochastic events

Reintroductions can be strongly affected by natural disturbances or stochastic events. Disturbances can be short-term, in the form of extreme weather events such as storms, floods, heat waves or fire, or can occur over longer time frames when extreme climatic conditions continue (IPCC, 2012). Individual events can cause high mortality within vulnerable populations, particularly when disturbances are extreme, occur in close succession or when resilience of populations is compromised (Saunders et al., 2011). Despite obvious direct impacts on populations, disturbances can form an integral part of ecosystem functioning, and can be necessary for the renewal of resources and persistence of many disturbance prone communities (Sousa, 1984). In these situations, changes to the natural disturbance regime (spatial and temporal dynamics of disturbances over time) can have long-term impacts on habitat structure and species survival (Turner, 2010).

Fire is a dominant disturbance in many ecosystems globally, and alterations to the natural frequency or intensity of fire can cause significant shifts in habitat condition and community composition (Bond & Keeley, 2005; Bowman et al., 2009; Kelly & Brotons, 2017). In systems like Australia where fire is a common disturbance process, it is often necessary for maintaining landscape heterogeneity and biodiversity (Bradstock et al., 2002; Woinarski & Recher, 1997). Fire influences landscape patterns of habitat and resource availability through redistributing important resources across both time and space (Ravi et al., 2009). As such, disturbance events such as fire can have strong impacts on reintroduced populations that may already be limited by suitable habitat extent and resource availability.

The flora and fauna of Australia have co-evolved with fire over millions of years, with the prevalence of fire and drought playing a significant role in shaping its fauna and flora (Bradstock et al., 2002; Gill, 1975). Following European settlement, the patterns of fire in the landscape have changed and many native habitats are poorly adapted to current fire regimes (Laurance et al., 2011; Mooney et al., 2011; Ward et al., 2001b). Land use changes, climate change, introduced species and continued destruction and degradation of ecosystems across the continent have increased pressure on native species, and may exacerbate the negative effects of inappropriate fire regimes (Laurance & Curran, 2008; Maron et al., 2015; Turner, 2010). Species with small populations that are already under pressure from extrinsic factors are hence more vulnerable to individual fire events (Shaffer, 1981). For example, highly fragmented landscapes may mean species in fire prone habitats, like the eastern bristlebird, are less able to recover from individual fire events due to their inability to disperse to unburnt areas (Baker, 2000). Conservation programmes attempting reintroductions in Australia therefore face the added paradoxical nature of fire in that it has direct effects on the mortality of individuals, but may also be a necessary action for restoring ecosystem functioning and species persistence. Suitability of an area of habitat for species persistence may therefore depend on whether it can provide adequate refuge during fire events as well as the resources immediately after fire for short- term survival. Habitat suitability in fire adapted ecosystems will also depend on whether the appropriate fire regimes needed to maintain habitat condition and critical resources are present, or can be re-instigated through active fire management. These considerations need to be based on complete ecological knowledge of the target species, as generalisations for fire management based on plant communities are often not suitable for promoting resources for fauna (Clarke, 2008; Woinarski et al., 2005).

ii. Predation

Invasive predators have had noteworthy impacts on many ecosystems worldwide (Moseby et al., 2015). In New Zealand, for example, the historical absence of native mammalian predators has produced native biota that are highly sensitive to introduced mustelids, feral cats (Felis catus) and rodents (Innes et al., 2010; Russell et al., 2015). In Australia, the introduced red fox (Vulpes vulpes) and feral cat have been largely responsible for the highest mammal extinction rates in the world (Johnson, 2006; Short & Smith, 1994).

Predation has been cited as one of the leading causes of reintroduction failures, with studies such as Moseby et al. (2011) highlighting the importance of predator free sanctuaries in determining reintroduction success. Reintroductions to areas where invasive predators have not been suitably controlled generally suffer high mortality rates (Combreau & Smith, 1998; McCallum et al., 1955; Short et al., 1992).

Reintroduction outcomes can be affected by both introduced and native predators. Changes in the distribution or abundance of native predators following habitat modification can affect their impact on normally resilient prey species. Naïve prey species from captive releases may also be more vulnerable to native predators. A reintroduction attempt of the endangered woma python (Aspidites ramsayi) was unsuccessful due to complete mortality of individuals by mulga snake (Pseudechis australis) predation, a previously unknown predator (Read et al., 2011). Read et al. (2011) suggest this high mortality may have been due to prey naivety or inability to find suitable refuge. Such failures demonstrate the importance of understanding potential predation risks and habitat-predator interactions within release habitat prior to reintroduction.

For ground dwelling species, appropriate understorey vegetation cover may be a crucial element needed for individual survival as they are often limited in their ability to disperse or escape predation. Predation rates may be higher in release sites where changes to vegetation cover or complexity affect a species' ability to find refuge and avoid detection (Christensen, 1980; McGregor et al., 2015; Vernes, 2000). Loss of important refuges such as logs and trees at release sites for the brown treecreeper meant longer flight times and resulted in higher predation risk to individuals (Bennett et al., 2013). Predation risk can also be exacerbated by disturbance events such as fire, and recent studies in Australia have revealed how fire scars (recently burnt areas) are highly attractive to feral predators, increasing predation in these areas (Doherty et al., 2015; McGregor et al., 2017; McGregor et al., 2014).

iii. Disease and pathogens

Until recently, consideration of disease has largely been neglected in reintroductions, with other aspects such as predator control usually seen as priority actions for managing

releases (Viggers et al., 1993). However, disease has the potential to have long lasting effects on reintroductions, particularly in small populations that have undergone extreme genetic "bottlenecks" and stress, which commonly accompany reintroduction attempts. An example of a bottleneck is the loss of genetic diversity that results from a reduction in population size. This loss can reduce the ability of individuals (and populations) to adapt to environmental change or resist diseases, have an impact on their reproductive output, and increase their vulnerability to stochastic events (Frankham et al., 2010). As a consequence, the suitability of habitat for reintroductions should be assessed in terms of the level of resilience they provide for these issues.

Incorporating disease into habitat suitability criteria has been especially important in amphibian reintroductions, where chytridiomycosis (caused by the Batrachochytrium dendrobatidis fungus) has caused infections in 42% of species globally (Olson et al., 2013) and been responsible for the rapid decline or extinction of 200 species (Skerratt et al., 2007). The fungus causes high mortality rates in the wild, and the devastating effects it has had on amphibian communities have led to its reputation as the worst recorded wildlife disease (Berger et al., 1999). Such emerging diseases, or changes in the infection rates or distribution of existing diseases, can have huge consequences for reintroductions, changing otherwise pristine habitat into highly unsuitable areas for population persistence (Berger et al., 1999). Another example where disease prevalence is a crucial factor in determining habitat suitability for reintroductions, is the Tasmanian devil (Sarcophilus harrisii). The species-specific, highly infectious devil facial tumour disease has caused a greater than 80 % decline in Tasmanian devil numbers in the last 20 years (Hawkins et al., 2006; Hogg et al., 2017; Lachish et al., 2007; McCallum et al., 2007), and is a prime example of the impact genetic bottlenecks have on susceptibility to pathogens (McCallum, 2008). In birds, the inadvertent introduction of avian malaria (Plasmodium relictum) to the Hawaiian Islands has been responsible for honeycreeper declines and extinction (Scott et al., 1986). The impact is expected to increase with climate change (Atkinson & LaPointe, 2009), highlighting that habitat suitability or resilience to such threats may change in the future. In each of these cases, determining habitat suitability for long-term persistence of reintroduced populations should ideally consider not only current pathogen distributions and behaviour, but also likely future disease dynamics and exacerbating factors.

Habitat connectivity

Alongside the availably of resources and protection from threats, the amount of habitat, along with its configuration in the landscape, are also important factors influencing the long-term persistence of species (Jackson & Fahrig, 2016; Root, 1998). Without a sufficient area of habitat, or a configuration that allows for reproduction and dispersal, small sub-populations are likely to be under more pressure from fluctuating resource availability, stochastic events or genetic bottlenecks (Fahrig & Merriam, 1994). Prior to reintroducing individuals, it must first be determined whether habitat patches are large enough to support viable populations or, if not, whether the connectivity between suitable habitat and other breeding populations is sufficient to maintain gene flow.

Many threatened species today have been reduced to small, isolated populations as a result of habitat loss (Saunders et al., 1991). For species with limited dispersal capability, fragmentation and isolation can have a greater impact on the long-term viability of populations and increase extinction risk (Fahrig & Merriam, 1985). A key concern of fragmentation, particularly in already small populations, is the loss of gene flow between sub-populations (Fernández et al., 2008). When populations are fragmented, the increased isolation of subsequent sub-populations has two primary genetic effects: increasing differentiation among sub-populations and decreasing genetic diversity within sub-populations (Keyghobadi et al., 2005; Templeton et al., 1990; Young et al., 1996 ). Genetic diversity is inherently connected to the resilience of populations, and so the loss of genetic diversity within threatened populations can have a substantial impact on long-term persistence of a species (Weeks et al., 2011). Much of the work addressing this threat in conservation programmes is done ex situ in captive breeding through management of breeding populations to promote genetic diversity (Weeks et al., 2015). Once reintroduction or translocation actions are an option, habitat suitability, especially for the long-term viability of populations, needs to account for connectivity and dispersal ability of individuals throughout the landscape to avoid in situ genetic depression.

Sites that lack either suitable habitat condition or appropriate configuration in the landscape are unlikely to be able to support populations in the long-term. If habitat condition is restored, but connectivity between isolated patches is not, there is a high chance that isolated populations will continue to decline and may become locally extinct from stochastic events or inbreeding effects. As the core goal of reintroductions is to create self-sustaining populations, the genetic diversity of the release group and the ability of release sites to be part of a functional meta-population will be important in determining long-term success (Weeks et al., 2011).

1.1.4. Habitat requirements of the eastern bristlebird

Habitat preference of the southern bristlebird is well described, and consists of a range of heath vegetation types that are dominated by a dense, thick shrub layer (Baker, 2000). They most often occur in coastal or montane heath habitats that range from dense heathy woodland to sedge or dense heath (Bain & McPhee, 2005; Baker, 1998b, 2000; Bramwell et al., 1992; Clarke & Bramwell, 1998). Today, most of the occupied heath habitats are protected to some extent in conservation reserves, so the main historical threats of habitat loss from coastal development and urbanisation are largely absent (D. Bain pers. comm.).

The southern bristlebird has been well studied, particularly in regards to translocation and post-fire response (Bain et al., 2008; Bain & McPhee, 2005; Baker, 1998b, 2000; Baker et al., 2012; Lindenmayer et al., 2009). Research on the southern bristlebird has helped identify the need for fire regimes consistent with the maintenance of areas of dense shrub cover, and the presence of refuge areas to allow populations to avoid exposure to the increased predation that follows high intensity fires in these habitats. Inappropriate fire is the main current threat to southern bristlebirds (Bain et al., 2008; Baker, 2000; Clarke & Bramwell, 1998). This work has provided an understanding that southern bristlebirds are highly sensitive to large fires, relying on dense, shrubby habitat for cover, and that fire exclusion is the best management action for encouraging the recovery and long-term persistence of southern bristlebirds (Baker, 2000). However, the understanding of habitat requirements for southern bristlebird is of limited assistance in conserving the northern bristlebird.

Northern bristlebird

The northern bristlebird population is highly fragmented, with discontinuous patches of habitat found within a rainforest and agricultural matrix and only holding a few individuals each. With such a reduced, highly fragmented population, and very few breeding individuals known to exist, the northern bristlebird is unlikely to persist without intensive intervention.

For intervention to be successful a strong understanding of the population's ecology and habitat requirements is needed. Ecological knowledge of the northern bristlebird has, until recently, been primarily derived from observation and monitoring, with limited scientific research to provide evidence for effective management decisions.

Unlike its southern counterpart, the northern bristlebird inhabits grassy eucalypt forest patches on the rainforest margin (Hartley & Kikkawa, 1994; Holmes, 1982; Lamb et al., 1993; Rohweder, 2000). Sandwiched within a landscape matrix of rainforest and cleared pastureland, the distribution of grassy eucalypt forest patches has declined, and remaining patches have been highly modified or degraded. Introduced weeds, clearing, grazing by livestock and changes to fire regimes are considered the main causes of decline and degradation of grassy forest habitat (Hartley & Kikkawa, 1994; Holmes, 1989; OEH, 2012; Rohweder, 2000, 2006; Wildsearch Environmental Services, 2016b). Despite this, there is still relatively little scientific understanding of northern bristlebird habitat requirements and the role that fire plays in maintaining habitat.

1.2. The problem

With such a small effective population size and a highly fragmented distribution of extant territories, the northern bristlebird is unlikely to persist without intervention. A pilot reintroduction attempt of northern bristlebirds from a former captive population (with relatively little preparation, precipitated by a forced closure of the facility), despite short

term survival of individuals (Stewart et al., 2009) was unsuccessful in the long-term, most likely due to the small number of birds released (four birds into two disjunct sites). The short term survival of birds indicated northern bristlebird are capable of being translocated into the wild (Stewart et al., 2009). To increase the chance of long-term persistence of reintroduction released birds and successfully aiding the long-term recovery of the northern bristlebird in the wild, an accurate understanding of the relative roles of vegetation structure, invertebrate prey and landscape context of open grassy sclerophyll forest habitat patches in determining habitat suitability for the northern bristlebird is vital. In addition, increasing our understanding of how these habitat characteristics are influenced by fire management will be important in determining what form of on-the-ground management is needed prior to reintroductions so that suitable habitat conditions for long- term persistence will be promoted.

1.3. Thesis aims & objectives

The overall aim of this thesis is to establish the habitat and disturbance factors associated with the persistence of northern bristlebirds in their grassy eucalypt forest habitat patches embedded within the Gondwanan rainforests of eastern Australia. To achieve this, the thesis examines the relationships between vegetation structure, patch size and context, invertebrate food abundance and fire history, and the persistence or local extinction of northern bristlebirds across their historic distribution (Fig 1.2).

CHAPTER 1 Introduction to research

HYPOTHESIS DEVELOPMENT

CHAPTER 2 Review of the ecology of the northern population of the Eastern bristlebird

Review of current information on ecology and habitat and identify knowledge gaps

HABITAT SUITABILITY THREATS

HABITAT STRUCTURE

CHAPTER 3 The importance of grassy patch size and FOOD RESOURCES structure for northern Eastern Bristlebird FIRE CHAPTER 4 persistence in a dynamic ecosystem Patterns of invertebrate food CHAPTER 5 availability and the persistence of an Are there differences in plant community Frequent fires are needed for insectivore on the brink composition, vegetation structure and landscape maintaining grassy forest habitat context among northern bristlebird sites? Is northern bristlebird decline related to for the northern bristlebirds reductions in prey availability? How have changes in fire regimes influenced northern bristlebird habitat?

CHAPTER 6 Conclusions, recovery recommendations and future directions

Figure 1.2. An overview of the thesis structure.

For the northern bristlebird, determining what habitat conditions are associated with their continued persistence will not only provide important information for their management, but also shed light on ecological changes that might be occurring within grassy eucalypt forest patches more broadly. This thesis aims to provide necessary ecological information that will contribute to on-the-ground management of habitat for northern bristlebird persistence and reintroduction attempts across southern Queensland and northern New South Wales.

The specific research questions addressed in this thesis are:

1 What is the current knowledge of northern bristlebird ecology? (Chapter 2)

2 Which structural characteristics of grassy eucalypt forest habitat are associated with northern bristlebird persistence? (Chapter 3) 3 Does food resource availability differ between habitat patches in which northern bristlebirds have persisted and those in which they have not? (Chapter 4) 4 How does fire history influence the distribution and suitability of habitat patches for northern bristlebird, and what fire regimes are most appropriate to maintain habitat for northern bristlebird? (Chapter 5) 5 Based on this new evidence, what is the best management strategy for conserving northern bristlebird habitat to improve the chance of long-term persistence? (Chapter 6)

1.4. Thesis structure

This thesis is organised as a series of manuscripts prepared for submission to, or accepted for publication by, scientific journals, and accompanied by an introduction and a synthesis. One chapter (3) is accepted by EMU-Austral Ornithology while chapters have been submitted and under review in Austral Ecology and the International Journal for Wildland Fire, respectively. This format necessitates some repetition, with introductory paragraphs for each chapter describing the study species, region and data collection methods. In some chapters, the same data has been utilised in a different analysis, hence the occurrence of some replication of method descriptions.

Chapter 1: Introduction and background.

The first chapter of this thesis provides a general introduction and overview of the problem, with the aims and research questions of the study. In Chapter 1, I review relevant literature on reintroduction biology, habitat suitability and resource requirements of threatened species, focussing on terrestrial ecosystems. I establish the overall aims of this thesis and provide a structural framework for the remaining chapters.

Chapter 2: Review of the ecology of the northern population of the eastern bristlebird

For northern bristlebird recovery to be successful and to maximise the chances of successful reintroduction efforts, it is important that current ecological and management information is properly reviewed and remaining uncertainties are correctly identified. If uncertainty is high, reintroduction may fail, which is a concern when reintroductions are costly to implement and the number of individuals available for reintroductions is limited. Identification of knowledge gaps can assist in the development of the most appropriate management strategy that optimises reintroduction success and species persistence. While the northern eastern bristlebird has been monitored extensively since the 1980s, much of the data collected exists in grey literature or as specialised knowledge by individuals within the Northern Working Group. Chapter 2 aims to provide an up to date synthesis of previous monitoring work and habitat assessments by systematically reviewing relevant grey literature and specialist knowledge from the Northern Working Group. I then use the results of this review to develop the hypotheses regarding northern bristlebird habitat and the role of fire that are tested in the following chapters.

Chapter 3: The importance of size and structure of grassy patches for northern eastern bristlebird persistence in a dynamic ecosystem.

While northern bristlebirds appear to prefer habitat patches with thick grassy understorey, the specific structural and floristic attributes, and habitat patch context that may contribute to persistence have not previously been systematically evaluated. Based on the potentially important factors identified during Chapter 2, Chapter 3 aims to test whether habitat patch context, structure, floristics or a combination of these have contributed to the persistence of northern bristlebirds. I achieve this by comparing the characteristics of, and contextual contributors to, the local habitat in territories where birds have persisted in and those where they have not. I use the results to examine and compare the relative importance of these habitat characteristics and contextual contributors as predictors of northern bristlebird presence.

Chapter 4: Availability of invertebrate resources and their contribution to persistence patterns of the northern population of the eastern bristlebird.

Globally, insectivorous birds appear to be particularly vulnerable to decline. The main hypothesis suggested is that this is due to declines in available food resources. With a predominantly insectivorous diet, it is possible that northern bristlebird decline may be associated with reduced food resources and that changes in environmental or vegetation characteristics of habitat patches may be responsible. This is particularly important to clarify prior to the reintroduction of individuals, as food resource limitations may reduce persistence and breeding success. Chapter 4 aims to determine whether northern bristlebird persistence is linked to the availability of invertebrate food resources across current and historic northern bristlebird habitat patches. I present an analysis of the abundance, community composition and nutritional value of invertebrate resources present within northern bristlebird habitat patches. I then used the results to test whether occupied northern bristlebird patches have different food availability than patches where birds are no longer present.

Chapter 5: Frequent fires are needed for maintaining grassy habitat for the northern eastern bristlebird

Changed fire regimes are one of the main mechanisms that have been suggested for the decline in the grassy habitat of the northern bristlebird and their own decline. Chapter 4 aims to determine whether and how the occurrence of fire in the grassy habitats of northern bristlebirds and its management impacts on their persistence. Changes in habitat extent between 1981 and 2009 were assessed in relation to fire history that occurred at each northern bristlebird site over this period. I then use this fire history data set to compare fire history between current and historically occupied habitat patches and thereby to determine how fire may influence northern bristlebird presence and habitat characteristics. I provide evidence of the importance of specific fire regimes in maintaining the extent and thick grassy structure of northern bristlebird habitat patches, and use the results to suggest fire frequencies most likely to maintain suitable habitat conditions for northern bristlebird persistence (as determined in Chapters 2 & 3).

Chapter 6: Synthesis & conclusion.

This final chapter presents a summary of my findings relating to each research question and discusses the contribution this thesis makes to the general ecological knowledge and management of the northern eastern bristlebird. It includes a detailed list of management recommendations specific to the northern bristlebird specific and based on the findings of Chapters 2 to 5. I also provide some discussion on the broader implications of this study to reintroduction biology, grassy forest conservation and invertebrate ecology. Finally, it provides a review of the limitations of this work and suggests future avenues of research that will help to maximise the potential for the success of northern bristlebird recovery.

Chapter 2

Review of the ecology of the northern population of the eastern bristlebird

Plate 1. Fire edge between recently burnt and unburnt grassy habitat in the Border Ranges.

2.1. Introduction

The eastern bristlebird (Dasyornis brachypterus) is a small, endemic, ground dwelling passerine that occupies low dense heathland and grassy vegetation in a severely fragmented distribution along the south-east coast of Australia (Baker, 2000; Lamb et al., 1993). It is listed as ‘Endangered’ under the Commonwealth Environment Protection and Biodiversity Conservation Act 1999 because of a steep decline in its distribution and abundance over the past 30 years. As a result of habitat loss and degradation, the once continuous distribution of the eastern bristlebird from Queensland to South Australia is now fragmented into three separate populations (Fig. 2.1).

QUEENSLAND Conondale Range Spicers Gap Sarabah Range Main Range Mt Barney NT Border Ranges Stretchers Track QLD Long creek Border Ranges Lismore Cedar Getters Pinnacle WA Razorback Green Pigeon Helmut Ridge SA Grafton Mt Burrell Dorrigo NSW Newport Middle Harbour Manly Parramatta VIC NEW SOUTH Port Hacking Bondi Wootton Wattamoola La Perouse Historical records WALES Madden Plains TAS Mt Kembla Northern population Bendeela Barren Grounds Central population Sydney Red rocks Budderoo Red rocks/Barren grounds Cambewarra Southern population Jervis Bay Bherwerre & Sassafras Beecroft Peninsulas BherwerrePenninsula Ulladulla Ulladulla VICTORIA

Nelson Nadgee/Mallacoota Marlo Lake Tyers Nadgee Mallacoota Yeerung River Bellbird creek Point Ricardo Clinton rocks Second Island Red River Bemm River Cape Conran Toorloo Arm

Figure 2.1. Historical and recent records of eastern bristlebird across south-east Australia. Adapted from Baker (1997), Lamb (1993) and Clarke & Bramwell (1998).

The total population size of the eastern bristlebird is estimated at 2500 individuals. The central population is the largest with a relatively stable population of roughly 2000 individuals (Bain et al., 2008; Baker, 1997; OEH, 2012). Genetic research by Roberts et al. (2011) found that the northern population had insufficient genetic differentiation from southern birds to retain the previously recognised subspecies D. brachypterus monoides.

Nevertheless, the northern population occupies substantially different habitat to its southern counterparts and there are several behavioural and plumage colouration differences between northern and southern birds (Chaffer, 1954; Holmes, 1989; OEH, 2012; Schodde & Mason, 1999). As such, the northern population is still considered a separate management unit that requires distinct management actions (OEH, 2012).

While the ecology and management needs of the southern populations have been well established, there is still little formal scientific research available on the habitat requirements and management needs specific to the northern bristlebird. Prior to this thesis, a total of 24 journal articles (based on a literature search using Web of Science and Scopus databases) about the eastern bristlebird have been published, and all but one focus on the southern New South Wales and Victorian populations. But, as already mentioned, the northern bristlebird inhabits high-elevation, grassy eucalypt forest patches within a largely rainforest-dominated landscape (Hartley & Kikkawa, 1994; Lamb et al., 1993; Wildsearch Environmental Services, 2016b). This habitat has vastly different vegetation composition and disturbance regimes to the heath and heathy woodlands inhabited by the central and southern populations, and thus habitat-specific research and population specific management is needed.

A National Recovery Team, with the objective of curating and helping guide implementation of a National Recovery Plan for the eastern bristlebird, was established in 1997 (Holmes, 1998). In recognition of the very different ecology of the northern population, a Northern Working Group was established from this meeting in 1998, with a northern specific draft recovery plan developed in 2001 and formalised with a Northern Working Group Business Plan in 2010 (Charley, 2010). At this time, the National Recovery Plan was reviewed, to include an update of priorities for management efforts to address and halt the decline of each bristlebird population (OEH, 2012).

The Northern Working Group includes experts from five government agencies, conservation and natural resource management non-government organisations, private consultants, three universities, a wildlife park (i.e. a conservation zoo), and private

landholders. A key objective outlined in the 2010 Business Plan was the re-establishment of a captive breeding program, with the aim of restoring the wild population to at least 156 birds and building a captive stock of 120-300 birds for release (Charley, 2010). In response to this objective, a captive breeding program commenced in 2014 that built upon the pilot program at David Fleay Wildlife Park from 2002-2009. The new program, along with increased efforts to implement more active habitat management including appropriate fire regimes and control of invasive weeds, has meant that reintroductions of captive-bred individuals into suitable habitat within historic northern bristlebird territories may soon be possible. However, little information has been available on habitat conditions needed for long-term success of reintroduction attempts.

For future reintroduction of northern bristlebirds to be successful it is vital that an accurate synthesis of known information and knowledge gaps is available. To accurately determine what gaps in knowledge are still present for northern bristlebird, we deemed it important to review the current existing grey literature to outline current knowns and unknowns of their ecology, distribution, habitat and threats. The overall objective of this chapter is to provide a preliminary investigation into our current understanding of eastern bristlebird ecology from which a conceptual model of habitat requirements and northern bristlebird persistence was developed and tested in Chapters 3-5 of this thesis.

Specifically, this chapter (Chapter 2) firstly provides a brief comparison of southern and northern bristlebird ecology, to emphasize the difference in knowledge and required management between the two. Secondly, I describe in more detail the current conservation status, distribution, habitat characteristics and threats acting on our focal northern population. Lastly I develop a conceptual model of suspected habitat dynamics of the northern bristlebird, with particular attention on the role of fire in maintaining or promoting appropriate habitat condition.

2.2. Methods

2.2.1. Literature review

In this review, I compile all known published and grey literature relating to the northern bristlebird. As this thesis centres on the northern population, the majority of this Chapter 26

will focus on this population, but a short review of southern bristlebird ecology has been included based on published research to highlight the differences between them. Grey literature included a range of technical and survey reports, recovery plans, husbandry manuals, historical habitat assessments and theses. Most of the literature consists of reports detailing regular monitoring of northern bristlebird territories, and habitat descriptions of these territories at their time of survey (the most recent being in 2016 (Pisanu et al., 2016)). Most of the documents are held by members of the Northern Working Group, while some were accessible online. Information used for this review was also derived from a range of experts, and synthesised to develop a then-current understanding of the factors influencing the northern bristlebird and its persistence.

2.2.2. Northern bristlebird distribution and population trends

I determined historical and current northern bristlebird distribution using the territory survey information of Holmes (1982) as the starting point, with territory occupation data coming from bi-annual monitoring of the territories since 1999 until 2016 (the most recent survey) by members of the Northern Working Group. The Northern Working Group has coordinated bi-annual surveys at the Queensland territories identified by Holmes (1982) and NSW territories identified by Sandpiper Ecological Surveys (2012) since 1999 using audio playback. This technique consists of an initial five-minute listening period, then a five minute period broadcasting northern bristlebird recordings, and finally a 20-25 minute listening period to detect any vocal responses from bristlebirds. If northern bristlebirds are detected within the initial listening phase, no calls are broadcast. Once birds respond, listening within that territory ceased. Surveys are conducted mostly during July-September when northern bristlebird begin to be more territorial at the start of the breeding season and the likelihood of response is high. At the sub-population level, once no bristlebirds were recorded in a territory for five years in a row, that territory was classified as ‘northern bristlebird absent’ and was no longer visited bi-annually, as recommended by Sandpiper Ecological Surveys (2005). Sporadic resurveys of these territories have been conducted to check for continued absence. In some cases, where individual territories have not been occupied for a long period, but still contain apparently suitable habitat and have neighbouring occupied territories, surveys were still conducted bi-annually. In cases where there was five years of absence and the habitat had changed significantly to the point

where it no longer contained grassy understorey (e.g. it turned into rainforest), surveys were abandoned altogether. As a result, survey effort across historical territories over time has varied.

Many northern bristlebird territories are situated close to each other, and were originally described as 1-2 ha in size by Holmes (1989) but appear to have expanded to 4-6 ha with lower bird densities within habitat patches (D. Charley pers. comm.). For the field work conducted in Chapters 3-5 of this thesis, it was necessary to group many of the smaller original territories into ‘sites’ because some individual territories were closely adjacent within habitat patches. As a consequence, this meant that the original 103 territories identified by Holmes were grouped into a total of 51 potential sites for survey in this thesis (Appendix 1). Of the 59 potential sites (original Holmes sites plus newer sites), five were historical sites in heath vegetation (located at Mt Barney, Mt Maroon and Surprise Rock) and another 12 no longer have any suitable grassy habitat remaining (e.g. Bald Knob, Green Pigeon and Cougal). These sites were excluded from survey due to the objectives of this thesis being on grassy habitat management actions for northern bristlebird recovery. The remaining four sites were excluded from survey due to access or time constraints.

In 2013, the NSW Office of Environment and Heritage began using a specially trained detection dog. This dog was trained to detect northern bristlebirds in the field with high confidence, and is capable of detecting northern bristlebird material (fresh, aged and diluted scat or feather samples) to a 2 m accuracy (L. Baker pers. comm.). The dog has since been used by the Northern Working Group since 2015 for ten surveys to confirm northern bristlebird absence at long-abandoned historical territories (Wildsearch Environmental Services, 2015b), confirm suspected presences at sites where birds have not been recorded recently (Wildsearch Environmental Services, 2015a), detect potential new territories in adjacent habitat to current territories (Wildsearch Environmental Services, 2016c) and locate nest sites within known breeding territories (Wildsearch Environmental Services, 2016a). Continued bi-annual monitoring of the same historical (Homes, 1989) and newer sites by the same experienced Northern Working Group members, along with the detection dog odour detection abilities allows for a high level of

confidence in presence/absence records for northern bristlebird across its historical range, despite its cryptic nature.

To determine overall population size over time, we collated all available monitoring reports into a single population database (Appendix 2). The northern bristlebird was first systematically surveyed in 1982 and 1984 (Holmes 1982; 1984), and again in 1988 (Holmes, 1989). In 1996 a follow-up survey was conducted (Holmes 1996). Since 1999 annual, and more recently biennial, surveys have been conducted (Sandpiper Environmental, 2005; Sandpiper Ecological Surveys 1999, 2000, 2002, 2003, 2006, 2007, 2011, 2012; 2014 D. Charley un-published data). For my review, survey data on number of individual birds present for each available monitoring report were collated to give a population estimate for that year. Data from the most recent 2015-16 bi-annual monitoring reports were used to estimate the current breeding population.

2.2.3. Habitat characteristics

A number of vegetation surveys and habitat assessments have been undertaken across the historical distribution of the northern bristlebird. In particular, Lamb et al. (1993) and Hartley and Kikkawa (1994) undertook detailed assessments of northern bristlebird habitat across both Queensland and New South Wales, which have been fundamental in describing northern bristlebird habitat. These documents are now over 25 years old, and anecdotal evidence by contractors have suggested considerable changes to habitat condition have occurred (D. Charley; D. Rohweder pers. comm.). More recent territory mapping and property management plans for landowners at occupied northern bristlebird sites have added to our knowledge of habitat condition and potential changes occurring within them (Rennison, 2016; Rohweder, 2000, 2006; Rohweder & Charley, 2008; Wildsearch Environmental Services, 2007a, 2007b). A report by Rennison (2016) mapped vegetation cover within a localised priority area (Border Ranges) between 1966 and 2009 to analyse changes in grassy habitat extent in the area. These reports have all been used in this review to provide an up-to-date assessment of northern bristlebird habitat and potential threats facing it.

2.3. The eastern bristlebird

2.3.1. Southern eastern bristlebird

The southern population, consists of an estimated 380 individuals (Bramwell et al., 1992; Clarke & Bramwell, 1998; OEH, 2012) and is found within Nadgee nature reserve in southern NSW and Croajingolong National Park in Victoria. The large central population occurs at inland sites at Barren Grounds and Red Rocks Nature reserves and coastal sites around Bherwerre Peninsula and Jervis Bay (Bain, 2006). The central population is the stronghold for eastern bristlebirds, with a stable population of > 2000 individuals (OEH, 2012, 2016). The southern and central populations are hereafter grouped for the purposes of this review (and referred to as the ‘southern bristlebird’) because of their very similar habitat and ecology.

The southern bristlebird occupies coastal or montane heathy habitats that are now mostly in conservation reserves. Heathy habitats can vary slightly from dense heathy woodland to sedge or dense heath (Baker, 2000, 2009; Bramwell et al., 1992; Clarke & Bramwell, 1998). Because much of their habitat is protected within either national parks or nature reserves, historical threats such as clearing for coastal development and urbanisation have been largely eliminated or were not substantial in these areas (Bain, 2006; Baker, 1997). Today, the key threat to the southern bristlebird in its remaining habitat is inappropriate fire regimes (Bain, 2006; Bain et al., 2008).

Inappropriate fire regimes in the past caused the loss of suitable habitat, either through temporary removal of vegetation following fires or through the modification of understorey structure to become unsuitable for southern bristlebirds (Fig. 2.2). High fire frequencies can reduce, and – if sustained – eliminate the dense thicket-forming shrubs needed by the southern bristlebird (e.g. Stage 3 → 4, Fig. 2.2) (Bain et al., 2008; Baker, 1997, 2000, 2009). After a fire event, because it takes time for the vegetation structure to develop a tall, thick shrub layer, the presence of neighbouring old growth heath refuges that have not been recently burnt is important for southern bristlebird recolonization (Baker, 2000). Because many of the southern bristlebird sub-populations are found in relatively small remnant patches that may be quite isolated, their ability to recover following fire events can be limited, particularly if the entire patch is burnt (Baker, 1997, 2000). The presence of

unburnt vegetation (Stage 3) allows southern bristlebirds to re-occupy territories quickly (within 2 years) following patchy fires (Bain et al., 2008; Lindenmayer et al., 2009). If no unburnt refuges remain (Stage 4), southern bristlebirds can take up to ten years to return to their pre-fire abundance in a patch (Baker, 1997; Woinarski & Recher, 1997).

High elevation heath - Barren Grounds

1 5

Transitional regrowth Heathland HABITABLE Dense, short canopy OPTIMAL

3 Fire Coastal heath - Jervis Bay Unburnt remnants HABITABLE

2 Extreme Heathland 4 Complete burn Wildfire No refuge Taller, open mid-stratum TEMPORARILY HABITABLE UNSUITABLE

Figure 2.2. Conceptual model of the habitat dynamics occurring within heath habitats of the southern and central eastern bristlebird populations in Victoria and New South Wales.

Southern bristlebirds have been well studied, and the fire-ecology of their habitat is relatively well known. In addition, because their habitats now largely occur in conservation reserves that have relatively natural fire regimes, the health and persistence of their habitat is largely assured. Because of this southern bristlebird population size is now relatively stable, although small. Fires that burn an entire area, leaving no neighbouring old growth habitat into which the birds can move, and from which they can readily return, is thought to be the major threat to southern birds (Bain et al., 2008; Baker, 2000). Thanks to these studies, we know that southern bristlebirds are fire sensitive, and that frequent fire is detrimental to their persistence and that fire exclusion has been a highly beneficial management strategy for southern bristlebirds. But for the northern bristlebird, which lives in grassy habitat, our story may be vastly different.

2.3.2. Northern eastern bristlebird

The northern population of the eastern bristlebird (hereafter the northern bristlebird) is isolated by more than 700 km from the central and southern populations (Fig. 1). Historically, this population was once found from Dorrigo (NSW), 300 km further south than the southern-most current birds, to the Conondale Ranges in the north (Holmes, 1989). The northern bristlebird population is highly fragmented across the landscape, and territories have become increasingly isolated in patches of grassy sclerophyll forest habitat within a rainforest and pasture matrix (Lamb et al., 1993; Rohweder, 2000).

Current status and decline

The northern bristlebird has an estimated population in the wild of fewer than 50 individuals (OEH, 2012; Sandpiper Ecological Surveys, 2012; Whitby, 2009; Wildsearch Environmental Services, 2016b). Collation of both official and unofficial monitoring reports (listed in Appendix 2) shows the population has declined by 80% (Fig. 2.3) from its largest surveyed size of 154 birds in 1989.

250

200

150

100

50 Northernbirsltebird population

0 1980 1985 1990 1995 2000 2005 2010 2015 2020 Year Figure 2.3. Population trends fitted using log-linear models for northern bristlebird since 1982 based on number of individual birds recorded during official surveys (squares, dotted trend line) and population estimates (circles, solid trend line) which includes additional observations from unofficial surveys. Based on monitoring reports and unpublished data between 1982 and 2016 (data from reports listed in Appendix 2).

During the first survey of northern bristlebird, Holmes (1982) identified 115 occupied territories, which after a follow up comprehensive survey (Holmes, 1989) had declined to 103 territories (with 12 territories becoming extinct). Only 51 of the 103 territories were occupied by paired individuals, while the rest only had single individuals recorded. By 1992, Lamb et al. (1993) found birds at only 32 of the 88 territories surveyed. Targeted monitoring of territories in NSW using audio playback and a specialised detection dog (since 2013) have revealed an increase in the number of adjacent pair territories at one location (D. Charley pers. comm.). As of July 2016, the population estimate for the northern bristlebird is 38 known individuals from at least 12 sites (Wildsearch Environmental Services, 2016b). Of these sites, only five (three in NSW, two in Qld) have confirmed breeding pairs (D. Charley pers. comm.). As a consequence, most currently occupied habitat patches (in which sites are found) today probably consist of only a few individual non-breeding birds, with nesting not observed at these locations. Low densities of birds in the wild spread across disjunct habitat patches means northern bristlebird territories are now highly isolated from each other and presumably have little to no dispersal and gene flow occurring (Roberts et al., 2011).

The small effective population size and highly fragmented distribution, mean that the northern bristlebird qualifies for ‘Critically Endangered’ under the IUCN red list criteria. It qualifies for this listing under criteria B2a = severely fragmented or known to exist at only a single location, B2b = continuing decline in (ii) area of occupancy, (iii) area/extent/habitat quality, (iv) number of locations/subpopulations, and (v) number of mature individuals, C2a = continuing decline in mature individuals and (i) subpopulation of < 50 individuals, and D = population size of < 250 mature individuals (Garnett et al., 2011; OEH, 2012). The northern bristlebird also likely qualifies for critically endangered status under criterion A2bc = reduction in population size of ≥ 80 % over the last 10 years or three generations based on abundance index and area of occupancy based on information given throughout this thesis.

The critical status of the northern bristlebird led to the development of a business plan by the Northern Working Group (2010). Key actions proposed in this plan were the control of

invasive weeds that were leading to habitat degradation, implementing prescribed burns to help maintain grass condition, and establishment of a captive breeding programme. In addition, the plan highlighted the need for additional research to quantify the habitat conditions associated with northern bristlebird persistence, particularly in terms of grass structure, food availability and identifying appropriate fire regimes for management. As a result of this plan, funding for this PhD was procured, and the captive breeding programme was re-established at Currumbin Wildlife Sanctuary. In addition funds have assisted landowners with prescribed ecological burns and weed control at sites where a need for fire was clear.

Habitat characteristics

Northern bristlebirds occur further inland and at higher altitudes then southern populations, and in grassy eucalypt forests close to rainforest. These areas are relatively cool and have high rainfall. They are now almost all located along the Queensland/New South Wales border (Holmes, 1989; Lamb et al., 1993). Historically a few northern bristlebird locations were in heath vegetation on high rocky areas, such as at Wollumbin/Mt Warning National Park or Surprise rock within Lamington National park, habitat that is more analogous to that of the southern populations. However, these heathland sites have not been occupied for almost 30 years, with the last recorded sightings in the late 80s and early 90s (Holmes, 1989, 1997).

Vegetation within northern bristlebird territories typically consists of a dense grassy understorey, with a sparse shrub layer and relatively open eucalypt canopy (Hartley & Kikkawa, 1994; Holmes, 1989, 1997; Lamb et al., 1993; Rohweder, 2000). The ground stratum within these habitats can be structurally diverse, with a range of grasses, vines, woody herbs and shrubs providing a heterogeneous grassy layer that provides shelter for northern bristlebird (Fig. 2). This vegetation provides important habitat for a range of other species in the Border Ranges region of NSW including 13 threatened species listed under the NSW Threatened Species Conservation Act (Table 2.1) (Sandpiper Ecological Surveys, 2000).

Table 2-1. Threatened species listed under the NSW Biodiversity Conservation Act (2016) recorded at northern eastern bristlebird locations. Adapted from Sandpiper Ecological Surveys (2000).

Common name Species name Threat status Source of record Pouched frog Assa darlington Vulnerable Incidental record Wompoo fruit-dove Ptilinopus magnificus Vulnerable NSW Wildlife Atlas Glossy black- Calyptorhynchus NSW Wildlife Atlas, incidental Vulnerable cockatoo lathami record NSW Wildlife Atlas, incidental Masked owl Tyto novaehollandiae Vulnerable record NSW Wildlife Atlas, incidental Sooty owl Tyto tenebricosa Vulnerable record NSW Wildlife Atlas, incidental Alberts lyrebird Menura alberti Vulnerable record Dasyornis NSW Wildlife Atlas, Targeted Eastern bristlebird Endangered brachypterus surveys NSW Wildlife Atlas, incidental White-eared monarch Carterornis leucotis Vulnerable record Barred cuckoo-shrike Coracina lineata Vulnerable NSW Wildlife Atlas Eastern tube-nosed Nyctimene robinsoni Vulnerable Incidental record bat Phascolarctos NSW Wildlife Atlas, incidental Koala Vulnerable cinereus record Hastings river mouse Pseudomys oralis Endangered Targeted surveys NSW Wildlife Atlas, incidental Yellow-bellied glider Petaurus australis Vulnerable record * Nationally listed under the EPBC Act

Lamb et al. (1993) undertook a detailed survey of the habitat across the northern bristlebird’s geographical range and determined that, like the southern bristlebird (Baker, 2000), understorey structure was key in determining habitat suitability. In particular ground cover, grass height and basal diameter of key tussock grass species were the best predictors (Lamb et al., 1993). Suitable habitat was found to be open forest with >75% ground cover, large tussocks (15-30 cm basal diameter) with heights in excess of 90cm, and with large areas of tussock grassland. Holmes (1989) also suggested that a minimum area of tussock grassland, especially S. leiocladum (Sorghum) was an important factor in breeding success.

Habitat can vary between and within northern bristlebird sites, with a range of understorey structural features observed (Holmes, 1989). The dominant grass species often varies, with more Imperata cylindrica (Blady grass), patches of thick tussock (i.e. Sorghum

leiocladum and Poa sieberiana/labillardieri), and non-grass species, such as ferns and vines, in wetter areas near rainforest margins (Fig. 2.4).

1 2 3

Greater habitat complexity is likely to help northern bristlebirds persist by providing a range of microhabitats. These may have a different probability of being burnt, be more hospitable during extreme weather events, or provide different food resources either spatially or temporally. For example, rainforest margins can often form a natural barrier to fire, and act as refuges during and following fire events (Holmes, 1989). Tree canopy cover heterogeneity also appears be an important influence of understorey grass structure, with canopy gaps allowing more light to the ground, and thicker tussock patches to persist.

Recent vegetation monitoring shows similar patterns in vegetation structure to those documented by Holmes (1989, 1997) and Lamb et al. (1993). Northern bristlebird territories are generally characterised by a thick grassy understorey, dominated largely by Themeda trianda, Imperata cylindrica or Poa species. The native tussock grasses are a particularly important feature of the grassy understorey for northern bristlebirds (Table 2).

Northern bristlebirds rely on large tussocks for providing shelter and nesting sites (Gubler et al., 2016; Hartley & Kikkawa, 1994; Holmes, 1989; Lamb et al., 1993; Rohweder, 2000, 2006; Young, 2003). Monitoring during the breeding season has shown a high preference for tussock patches with an open tree canopy for nesting. Captive bred individuals also have a high preference for open sunny areas, with captive nests generally constructed in the more open parts of enclosures (A. Beutel pers. comm.). Habitat with more extensive mature tussock patches (Holmes, 1989) and high grass diversity (D. Charley pers. comm.) is thought to support a higher number of northern bristlebird territories. Field observations have suggested that complexity within the grassy understorey is a crucial element in northern bristlebird habitat, and at some sites, territories in which grass structure has become uniform and extremely dense (often observed with I. cylindrica and T. triandra) have been abandoned (D. Charley pers. comm.).

Following the initial surveys in 1989, Holmes (1997) reported a decline in the proportion of extant northern bristlebird territories that had notable S. leiocladum patches, as well as a decline in the overall abundance of S. leiocladum within territories. Photo points taken in 1985 and 2013 at four sites at Richmond Gap confirm this change in understorey (D. Charley pers. comm.), with significant reduction in tussock cover in the understorey and increasing shrub and weed invasion (Fig. 2.5). Ground cover data from the 1989 and 2013 surveys are consistent with a postulated decline in S. leiocladum abundance (Table 2.1).

Table 2-2. Common plant species found within northern bristlebird territories. Figures are percent cover in territories originally identified by Holmes (1989). More recent figures are from Lamb (1993) and Stone (2013). 1993 and 2013 surveys focussed on changes in understorey vegetation therefore no tree data is available.

Common name Scientific name 1989 1993 2016 Trees & shrubs Sydney Blue Gum Eucalyptus saligna 55 Forest She-oak Allocasuarina torulosa 55 New England blackbutt Eucalyptus andrewsii 50 Austral Grasstree Xanthorrhoea glauca 50 Acacia Blackwood/Hickory 47 melanoxylon/implexa Pink Bloodwood Eucalyptus approximans 45 Tallowwood Eucalyptus microcorys 42 Grey Gum Eucalyptus punctata 40 Brush Box Lophostemon confertus 35 Green Wattle Acacia irrorata 29 Rough Barked Apple Angophora floribunda 24 Broad leaved Apple Angophora subvelutina 23 Forest Red Gum Eucalyptus tereticornis 22 Manna Gum Eucalyptus viminalis 19 Ground covers Blady grass Imperata cylindrica 90 91 94 Poa Snow grass 86 95 82 sieberiana/labillardieri Kangaroo grass Themeda triandra 74 60 76 Bracken fern Pteridium esculentum 74 45 38 Wild Sorghum Sorghum leiocladum 69 72 43 Spiny-head mat rush Lomandra longifolia 30 26 51 Prickly rasp fern Doodia aspera 3 - 16

There is some evidence that Pteridium esculentum (bracken fern) has decreased across sites. While this is anecdotal, bracken fern is commonly associated with fire, and is often common at sites where regular fire intervals have been retained (Spencer & Baxter, 2006; Tolhurst & Turvey, 1992). Overall, the grassy layer of occupied sites appears to have high structural complexity.

1985 1993 2013

Breeding

The presence of suitable nesting sites is an important consideration for ground-dwelling species that are highly vulnerable to predation and exposure, particularly in grassy systems (Dobson & Lyles, 2000; Nicoll et al., 2003; Reardon et al., 2012). Nesting sites need to provide adequate cover and protection against environmental conditions and predation, and provide food resources during breeding (Jones et al., 1995). Nesting sites that provide such benefits are likely to be a key limiting resource for breeding activity in the northern bristlebird (Lamb et al., 1993). The northern bristlebird breeding season is relatively long, and is generally considered to occur from August to February, peaking around October-November (OEH, 2012; Young, 2003). Evidence from field surveys and nest searches in NSW during the 2014-2016 surveys, however, suggest that breeding may be largely opportunistic, with northern bristlebirds capable of breeding throughout the year given the right conditions (D. Charley pers. comm.). Breeding of captive birds at Currumbin Wildlife Sanctuary has also shown this. As a result, the breeding season has been redefined as early as July through to March, with only a short non-breeding period between May/ June (Gubler et al., 2016). Breeding territories often have high tussock cover, and nests are predominantly constructed at the base of large tussock clumps, 10- 45cm off the ground (Gubler et al., 2016; Holmes, 1989). Northern bristlebirds are highly sensitive to disturbance at the nest, and nest abandonment is common after disturbance in the wild (Chaffer, 1954; Hartley & Kikkawa, 1994). This has been reported in response to

human disturbance and grazing, but it seems likely that this could also occur in response to predators or fire.

The northern bristlebird appears to be experiencing poor nesting and breeding success in the wild, with only five breeding pairs identified. While breeding success may be being impacted by the isolated nature of territories and reduced genetic diversity, the evidence above also suggests that the reduced abundance of tussocks clumps may also be playing a role.

Threats

Nationally, habitat fragmentation and degradation are considered the main threats acting on the eastern bristlebird (OEH, 2012). This has occurred through the disruption of natural fire regimes and urban land development, the later particularly important for the southern bristlebird which occupies prime coastal land (OEH, 2012).

Much of the conservation effort for the eastern bristlebird has focused on habitat and fire related threats, and as a result little is known about the impact of other threats such as predation. As ground-dwelling birds with poor flight capabilities, predation is likely to influence them, especially during breeding when nests are constructed low to the ground. The incubation period at 2-3 weeks (21-23 days in captivity) is relatively long (Gubler et al., 2016; Hartley & Kikkawa, 1994). Likely native predators include snakes, birds of prey, goannas, quolls, currawongs and honeyeaters, while introduced predators likely to prey on eastern bristlebirds include cats, foxes and pigs (OEH, 2012). The threat of habitat loss from urban development has largely been reduced for the southern bristlebird because all locations are on protected land, so the main management actions are protection against high intensity fires and some predator control (D. Bain pers. comms). Predation of northern bristlebirds is currently viewed as a lower priority than the restoration of habitat and appropriate fire regimes. Cat and fox densities are believed to be naturally low within occupied northern bristlebird territories. If appropriate fire regimes and the associated resources required for successful breeding are successfully restored, and northern bristlebird recovery is still not occurring, it may suggest that predation pressures are higher

than currently thought. If this is the case, more integrated predator and fire management may be required.

Fire

Fire is a common occurrence in eastern bristlebird habitat, with both southern and northern bristlebird habitats subject to periodic fire (Baker, 1997; Holmes, 1989; Lamb et al., 1993). The fire regimes appropriate for maintaining the different habitats, however, differ considerably.

The grassy understorey in northern bristlebird habitat means it is highly flammable but probably also dependent on fire for continued persistence against rainforest and weed encroachment (Lamb et al., 1993; Rohweder, 2006). Research on similar grassy vegetation types elsewhere in Australia shows that infrequent fire allows the encroachment of woody shrubs from adjacent vegetation types (Butler et al., 2014; Fensham & Fairfax, 1996, 2006; Hoffmann et al., 2012b; Moravek et al., 2013). For species that rely on grassy habitats, such as the northern bristlebird, this is of serious concern, as a shift in land use practices since European settlement has largely led to decreased fire within these ecosystems (Bowman & Prior, 2004; Bradstock et al., 2002; Kershaw et al., 2002). In addition, the lower slopes of many northern bristlebird sites have been affected by livestock grazing and introduced weeds such as Lantana camara (lantana) and Ageratina adenophora (crofton weed), and thus have had their structural integrity and flammability affected.

Originally, Holmes (1989) suggested frequent fires (every 5 years) in northern bristlebird habitat were reducing abundance of large tussocks necessary for breeding. As such the proposed management strategy was for prescribed fire intervals of 10-20 years to allow the grass understorey to flourish and reduce weed encroachment. Since this original report, it has become increasingly evident that such long intervals are likely to have a negative impact on northern bristlebird habitat (OEH, 2012; Rohweder & Charley, 2008; Sandpiper Ecological Surveys, 2012; Wildsearch Environmental Services, 2007b, 2016b). Rohweder (2006) proposed fire intervals of 5-10 years. Hartley and Kikkawa (1994) found that healthy grass habitat could recover to suitable foraging quality for northern bristlebirds

by six months post fire, and may be suitable for breeding by two years post fire (Rohweder, 2006). Birds have also been recorded in territories at six months post fire, but whether this was for breeding or just foraging is not known (D. Charley pers. comm.). Sites 4-5 years post fire seem to experience a decline in the structural condition of grassy understorey, and as such recent recommendations of 3-6 year fire intervals have been proposed (Watson & Tasker, 2013). So, while some aspects of what might be an appropriate frequency of fire to maintaining northern bristlebird habitat have been postulated, the details have not been clearly articulated in previous studies nor tested with extensive historical data.

While fire is needed for long term habitat maintenance, fire events may represent a high mortality risk for small and isolated remnant populations. Unburnt refuges are known to be important for the southern populations during the short term resource bottleneck and exposure to predation that arises immediately following fire (Bain et al., 2008; Baker, 1997; Lindenmayer et al., 2009; Pyke et al., 1995). So the frequency, intensity and extent of fire will be key to determining the structure and suitability of vegetation for both short term survival and long-term persistence (OEH, 2012). To manage northern bristlebird habitat, it is therefore imperative that the relationship between habitat quality and fire is correctly identified.

1.3.3. Potential habitat dynamics of the northern bristlebird

The existing work on northern bristlebirds has demonstrated the importance of fire in maintaining their habitat. Using the habitat information in the existing reports described above, I present a conceptual model of the potential habitat dynamics occurring within northern bristlebird habitat (Fig. 2.6).

1 3

Too frequent fire

Homogenous grass Eucalypt forest understory Tall grassy understory MARGINAL OPTIMAL 2

Appropriate fire regime Transitional threshold Eucalypt forest 4 5 Sparse grassy understory MARGINAL

No fire Rainforest Shrubby understorey UNSUITABLE TRANSITIONAL

Figure 2.6. Conceptual model of potential habitat dynamics occurring within northern bristlebird habitat based on expert knowledge and monitoring data.

Presently, optimal habitat that promotes successful breeding is thought to be grassy eucalypt forest with a thick, diverse grassy understorey (Stage 1). Understorey diversity is likely to be an important component that allows resources to persist that are needed for northern bristlebird survival following fire events. A critical question that emerges from this is:

1. Are there specific habitat features in terms of vegetation, landscape context or food availability that can be attributed to optimal sites?

Vegetation grows quickly in northern bristlebird habitat thanks to the warm summer temperatures, high rainfall and fertile soils. This environment means that there is a fast

turnover from grassy understorey to shrubby understorey if fires do not occur frequently (Tasker, 2017; Tasker & Watson, 2016, 2017). Change from optimal to marginal understorey structure is most likely to be associated with reduction in the tall, thick grass layer as grasses senesce, have an increasing amount of dead matter, and tussocks collapse (Stage 2). Compact grasses may reduce opportunity for vulnerable ground foraging species to move freely beneath the protection of grass cover (Whittingham & Devereux, 2008; Whittingham & Evans, 2004). This habitat stage, while likely to remain suitable for northern bristlebird foraging and individual survival, may be less suitable for breeding activities that require large, tussock patches in canopy gaps (D. Charley pers.comm.). At this stage, fire events help maintain a ‘fire trap’ for young shrubs and saplings and allow for the regeneration of the optimal thick grass cover (Stage 2 → 1) (Tasker & Watson, 2016, 2017).

Fire that is too frequent may be responsible for the loss of grass structural diversity observed at some northern bristlebird sites (Stage 3). With frequent fire, many sites have become dominated by only one or two grass species (e.g. T. triandra or I. cylindrica), and have lost the structural complexity that is important for northern bristlebird persistence (D. Charley pers. comm.). Stage 3 habitat is likely to still support northern bristlebird foraging and survival, but the loss of tussocks for breeding is detrimental for long term persistence (Lamb, 1993). I propose that habitat stages 1-3 are likely to be the range of habitats that are most beneficial for long-term persistence. While both ‘marginal’ habitat stages (2 and 3) are sub-optimal for breeding they are still capable of supporting northern bristlebirds. A mosaic of all 3 stages might produce high vegetation complexity, providing for foraging, breeding and temporary spatial refuge from fire events (Lamb, 1993; D. Charley pers. comm.). Persistence of optimal tussock habitat in some capacity could also maximise availability of a seed source of grasses.

This leads to the question:

2. How does resource availability and vegetation structure within grassy forest habitat patches change in relation to fire history?

The dashed line in Figure 2.6 indicates what I propose to be a transitional threshold for northern bristlebird occupancy, in which appropriate habitat transitions into unsuitable rainforest habitat. Over time the absence of fire allows seedlings and saplings to grow beyond the fire trap and begin to shade out the grass layer, eliminating large tussock patches that are dependent on high light conditions (Stage 4). Maintenance of a fire trap is especially important for species that rapidly become reproductively mature or competitively dominant as once trees are capable of producing seed, shrub encroachment and associated shading of the grass understory can accelerate quickly (Stage 4 → 5) (P. Watson pers. comm.). If fire is not introduced within a critical fire interval, i.e. before invasive plants are capable of reproducing, then any subsequent fires will be less able to supress their abundance. Field observations by members of the Northern Working Group are consistent with this idea, with many sites appearing to have degraded at an exponential fast rate over the past 30 years, reportedly where fire has been less frequent (D. Charley, D. Rohweder, pers. comm.). In some cases, sites have completely changed into rainforest with no sign of the once dominant grassy understorey. Some evidence suggests this complete habitat transition can occur in as little as 7-10 years when no fire is present (D. Charley pers. comm., Fig. 2.5).

A third critical question is therefore:

3. Is the loss of fire in grassy sclerophyll forest habitat responsible for northern bristlebird absence, and – if so – what fire regime is needed for their persistence?

The proposed conceptual model, based on expert knowledge and previous studies of northern bristlebird habitat by the Northern Working Group suggests that fire may be important driver of northern bristlebird persistence through habitat modification and potential loss of important resources. There are still significant gaps in our knowledge of the attributes that constitute suitable habitat, and the specific role fire has in regulating resources and habitat structure. Implementing fire management prematurely could have more negative outcomes on northern bristlebird recovery than intended. Through this review I have identified key questions that need to be addressed, and form the foundations of the following chapters in this thesis. During the following chapters I will answer the questions:

1. Are there specific habitat features in terms of vegetation, landscape context or food availability that can be attributed to optimal sites? 2. How does resource availability and vegetation structure within grassy forest habitat patches change in relation to fire history? 3. Is the loss of fire in grassy sclerophyll forest habitat correlated with northern bristlebird absence, and – if so - what fire regime is needed for their persistence? 4. Based on evidence present in this thesis, what is the best management strategy for conserving northern bristlebird habitat to improve the chance of long-term persistence?

Chapter 3

The importance of grassy patch size and structure for northern Eastern Bristlebird persistence in a dynamic ecosystem

Emu - Austral Ornithology, 1-12. doi:10.1080/01584197.2018.1425628

Plate 2. Thick grassy sclerophyll forest habitat of an occupied northern bristlebird site in the Border Ranges.

3.1. Abstract

Reintroductions are a critical tool for threatened species conservation, and their success will depend on correctly identifying the key habitat requirements needed for persistence. The critically endangered, isolated northern population of the eastern bristlebird has declined to an estimated 40 individuals. A successful captive breeding programme means future reintroductions are likely; however, knowledge of fine scale habitat requirements associated with persistence is limited. To inform reintroduction efforts, we compared habitat attributes of currently occupied and historically occupied habitat patches. Persistence was highly dependent on the extent of contiguous grassy habitat, with larger patches more likely to contain bristlebirds. This association was contingent on grass structure, with less chance of bristlebirds occurring at large sites lacking a tall, thick grassy understorey. Topographic and environmental heterogeneity within large habitat patches appears likely to allow unburnt habitat to persist following fire, increase the temporal availability of prey and allow persistence of a population. For a largely ground-dwelling species, the presence of tall, thick grasses is expected to provide important shelter for foraging and nesting activities. Use of appropriate fire to maintain large contiguous patches with thick, tall grassy ground layer will be critical for the continued persistence and successful reintroduction of the northern eastern bristlebird.

3.2. Introduction

Increased pressure from habitat degradation, habitat loss and introduced species has meant that many severely threatened species now require human intervention for long- term persistence. Intensive intervention may be required when populations are extremely small and highly fragmented, especially where recolonization rates are low, and the ability to form a self-sustaining population has been lost (Armstrong & McLean, 1995). In such cases, ex-situ conservation efforts like captive-breeding programmes and reintroductions are needed to boost populations. Strategic manipulation of these conservation tools can help increase genetic diversity of bottlenecked populations, and ultimately assist in species’ recovery (Jamieson et al., 2008).

Translocation and reintroduction have become accepted management actions for the conservation of small populations (Griffith et al., 1989; Seddon et al., 2007; Tear et al., 1993). Reintroduction (intentional movement of individuals to parts of a species’ historic range from which it has been lost) has a strong history in Australasia, and has become a core part of conservation strategies for a range of threatened species, particularly for birds in New Zealand and marsupials in Australia (Armstrong et al., 2015). Both countries have seen varying levels of success, but reintroductions have contributed to the recovery of some high profile species such as the Chatham Islands black robin (Petroica traversi), South Island saddleback (Philesturnus carunculatus), numbat (Myrmecobius fasciatus), burrowing bettong (Bettongia lesueur), and the southern population of the eastern bristlebird (Baker et al., 2012; Butler & Merton, 1992; Friend & Thomas, 1995; Jones & Merton, 2012; Short & Turner, 2000). Although these reintroductions have been successful, many others have failed (Bennett et al., 2012; Moseby et al., 2011; Short et al., 1992).

Failed reintroductions often result from inappropriate management of habitat conditions or threats prior to release, or from incomplete knowledge of species ecological requirements (Armstrong & Seddon, 2008; Griffith et al., 1989; Seddon, 2010). Reintroduction failure can be highly detrimental to species recovery, with many programmes limited in source population sizes and resources for such intensive management actions (Andelman & Fagan, 2000; Lindenmayer et al., 2002; Ricciardi & Simberloff, 2009; Simberloff, 1998; Tear et al., 1993). As a consequence, a key question for conservation managers to consider prior to reintroduction attempts is: what habitat conditions are needed for persistence, and are they present at release sites (Armstrong & Seddon, 2008)?

Establishing whether release sites have the necessary conditions for successful reintroduction is hard enough where habitat is relatively stable through time, but can be further complicated in dynamic systems. In dynamic ecosystems, natural and often essential disturbance events can have direct and indirect impacts on a species. These impacts can include direct mortality from the disturbance event itself or the rapid creation or loss of suitable habitat patches (Reigada et al., 2015; Turner, 2010). Many species require specific structural or compositional habitat attributes that result from particular

disturbance regimes (Fuhlendorf et al., 2006; Monamy & Fox, 2000). Ensuring that appropriate habitat conditions are available through space and through time can therefore be crucial to successful reintroduction of such species.

The eastern bristlebird (Dasyornis brachypterus) is an endemic passerine of south-eastern Australia that is classified as nationally endangered under the Environment Protection and Biodiversity Conservation Act (Bain et al., 2008). It is a small, cryptic bird with limited flying ability (Hartley & Kikkawa, 1994). Its distribution is severely fragmented between larger southern populations in NSW and Victoria and a small northern population, 700 km away along the QLD/NSW border (OEH, 2012). Currently, the northern and southern populations are treated as separate management units because of their very different habitat requirements (OEH, 2012). A recent genetic study did not find sufficient basis to support sub-species distinctions (Roberts et al., 2011). The southern populations occupy dense coastal and montane heath and heathy woodland vegetation, and thanks to appropriate habitat conservation and management are relatively stable (Baker, 1997, 1998b, 2009; Baker et al., 2012). Much of the scientific research to date on eastern bristlebirds has focussed on these southern populations. In contrast, ecological knowledge of the northern population, which occurs only in open grassy sclerophyll forest and woodland, has until recently been derived primarily from observation and monitoring, with limited scientific research done.

Since monitoring began in the late 1980s, the northern eastern bristlebird population (hereafter ‘northern bristlebird’) has declined by 80% (OEH, 2012; Wildsearch Environmental Services, 2016b). This decline is thought to be associated largely with reduced disturbance and shrub invasion in its grassy sclerophyll forest habitat (OEH, 2012; Rohweder, 2000, 2006). The 2016 annual census estimated the current wild population to be just 40 individuals, of which only five breeding pairs were confirmed (Wildsearch Environmental Services, 2016b). This tiny effective population size qualifies the northern population for Critically Endangered status under the IUCN (OEH, 2012). The extremely small and fragmented nature of the northern population means continued persistence is highly dependent on management intervention. One such intervention is a captive breeding programme, and recent successes in this programme mean

reintroductions back into the wild will soon be possible. To avoid reintroduction failure, good understanding of the habitat conditions needed for northern bristlebird persistence in the wild is urgently needed. Assessment of these conditions will aid in identifying the most suitable areas for reintroduction and determining appropriate management to maintain appropriate habitat.

Northern bristlebirds occupy isolated patches of grassy eucalypt forest in the mountain ranges along the QLD-NSW border. Currently, birds remain in just 16 of the known 38 historical forest patches. This persistence may be a consequence of specific habitat features within a patch, or the size or context of the patch itself. If historical habitat patches (where birds no longer occur) lack important attributes, future reintroductions to these sites may fail. As a ground-dwelling species, structural attributes of vegetation in the understorey (particularly the ground stratum) are thought to be particularly important for providing shelter. Diminished fire frequency in historical sites may have led to changes in vegetation structure and plant species composition that have reduced habitat suitability. In addition, the dynamic nature of grassy sclerophyll forests mean that northern bristlebird persistence may also be influenced by the broader landscape context of habitat patches. Here, we examine whether northern bristlebird persistence across its historical distribution is related to habitat structural or site contextual attributes. Specifically we test:

 Whether there are differences in the plant community composition between currently occupied and historically occupied sites;  Whether northern bristlebird presence at sites is associated with specific structural (grass height, cover) or landscape context (grassy habitat extent, distance to refuges) attributes; and  Whether northern bristlebird presence can be predicted accurately based on a combination of grassy habitat patch attributes.

3.3. Methods

3.3.1. Study Area

Fieldwork was conducted at all known historical and currently occupied northern bristlebird territory locations (hereafter, ‘sites’) that occur within grassy sclerophyll forest patches and that are embedded in a wet forest-rainforest matrix. While a few territories were known historically from montane heath habitat these are all now long abandoned, and thus we had no current heath sites against which to assess change. We therefore excluded them from this study. Grassy sclerophyll forest patches are highly dynamic, and occupy an important transitional zone between rainforest and dry open woodland in high rainfall areas of Australia (Campbell & Clarke, 2006). Such patches are important habitat for the northern bristlebird, northern bettong (Bettongia tropica) and hastings river mouse (Pseudomys oralis) (Lamb et al., 1993; Pyke & Read 2002; Vernes & Pope, 2001). Recently, encroachment of rainforest and weeds due to fewer fires have reduced the extent and condition of this habitat (Gooden et al., 2009; Harrington & Sanderson, 1994; Rennison, 2016; Williams, 2000).

Northern bristlebird sites are scattered, with clusters found in and around Conondale, Main Range and Lamington National Parks in QLD, and the Border Ranges National Park in NSW (Figure 3.1). The region’s climate is subtropical, with wet, warm summers and dry winters. Rainfall averages 947 mm/year, mostly during the summer, and air temperatures average 32 °C in summer and 22-26 °C in winter (obtained from Bureau of Meteorology, 2014). Sites are mostly on upper slopes or exposed ridgelines of mountainous topography, where rainforest gives way to more open woodland habitat. Much of the eucalypt woodland on lower slopes has been cleared for agriculture, with rainforest the dominant vegetation at higher altitudes. As a consequence, northern bristlebird habitat patches are now often bordered by unsuitable cleared pastoral land and rainforest.

Conondale Range

QUEENSLAND

Main Range NT Border Ranges QLD WA SA NSW NEW SOUTH Brisbane VIC WALES Historical EBB records Northern population TAS Central population Red rocks/Barren grounds Southern population Bherwerre & Beecroft Peninsulas Main Range VICTORIA Lamington Mt Barney Nadgee/Mallacoota Border Ranges

3.3.2. Habitat assessments

Vegetation surveys were conducted across all the 38 historic and currently occupied northern bristlebird sites, of which 16 are currently occupied. Sites were based on original territories defined by Holmes (1982) and home range estimates of newer contemporary northern bristlebird locations (Wildsearch Environmental Services, 2016b). These sites are highly clustered within regions, so a random sampling approach was used to locate four to six 20 × 20 m vegetation plots (depending on size of grassy patch) within each site. Vegetation plots were placed randomly within 200 m of recorded northern bristlebird locations within each grassy habitat patch, which is clearly distinguishable from the surrounding dense rainforest. To assess habitat condition, structural variables were measured within the ground vegetation (0 – 1 m above ground), shrub (1-3 m) and canopy (>3 m) strata.

Grass height was measured along the two central axes of the 20 × 20 m plot, at 50 cm intervals. Measurements were taken by dropping a 20cm diameter disc (20g) down a measuring pole to determine the height of the grass bulk that provides cover for northern

bristlebirds. These were then averaged to determine mean grass height for each vegetation plot. All plant species within each 10 x 10 m subplot were identified to species or genus level (when species could not be identified) and cover of each species within subplots was estimated. Overall cover for each identified species was then estimated for the ground (<1 m), shrub (1-3 m) and canopy (>3 m) vegetation strata of each subplot and then averaged across the entire plot. Cover estimates were visually assessed using methods outlined by Walker and Hopkins (1990). Cover estimates were used as counting individuals (abundance), particularly for graminoids, was not feasible. Cover estimates for each graminoid species were also summed to represent the vertical structure of grass cover within the 0-1 m strata (total summed grass cover).

3.3.3. Presence/absence data

Northern bristlebird presence and absence within habitat patches was determined using data from annual monitoring of all known historic territories (identified by Holmes 1982) which began in 1999. The monitoring consists of an initial two minute listening period, a five minute broadcasting period of northern bristlebird recordings, and a 20-25 minute listening period to detect bird response (Wildsearch Environmental Services 2016). This standardised annual monitoring by a small number of experienced observers allows for a high degree of confidence in the consistency and repeatability of surveys, which is unusual for a small, cryptic threatened species. In addition, a specialised detector dog with the NSW Office of Environment and Heritage was trained in 2013 to detect northern bristlebirds in the field. This detector dog has assisted in confirming presence and absence at historic territories and with locating birds for the captive breeding program (L. Baker pers. comm.).

3.3.4. Spatial characteristics

The extent of suitable grassy habitat within each site was mapped by visually delineating grassy patches of northern bristlebird sites using ADS40 aerial imagery (taken in 2009) with a resolution of 5 m. Rainforest has a distinct canopy appearance, so rainforest boundaries are easily distinguished by a stark contrast in green tone between eucalypt and rainforest canopy, denser canopy foliage and no canopy gaps. For modified

(disturbed) boundaries, a clear contrast between cleared eucalypt pasture and intact eucalypt forest canopy can be seen. Due to the high resolution photography used, grazing in areas with some intact canopy can still be visually differentiated from non-grazed grassy habitat, based on sparse canopy cover and understorey colour and texture. eastern bristlebird habitat was defined as areas with a grassy understorey in which at least one territory (hereafter, site) has been recorded. This definition was an amalgamation of two detailed habitat classes, ‘primary’ and ‘secondary’ habitat, defined by Rennison (2016) as part of a northern bristlebird habitat suitability project for NSW (Table 3.1).

Table 3-1. Description of northern bristlebird grassy habitat and non-habitat used to define habitat patches. Adapted from Rennison (2016).

Habitat type Understorey structure Canopy structure Grassy habitat Areas of woodland or open woodland Dominance of native with relatively unmodified grassy tussock grasses; Native Woodland (10-30% understorey; Areas of woodland or grasses with shrubs; PFC*); Open open woodland with evidence of Sparse dry shrubs with woodland (5-10% disturbance or shrubby elements but grasses; Native/Non- PFC) remain relatively intact; Close proximity native Managed Pasture to rainforest Non-habitat Areas of open forest dominated by a shrubby understorey. Closed forest. Shrub dominated; Moist Woodland or open woodland where Open Forest (30- forest shrub species; disturbance or management practices 70% PFC); Closed Weed dominated; Grazed have removed native tussock grasses. forest (70-100% or significantly modified Areas cleared of native vegetation. PFC) grasses Areas not in close proximity (1km) of rainforest vegetation. * PFC = permanent forest cover

Habitat patches were mapped as the total continuous extent of grassy habitat contiguous with the location in which a northern bristlebird has been recorded. In some cases, several separate sites (territories) were located in a single large continuous habitat patch. When this occurred, the same (i.e. one single figure) habitat patch area was used, as this total extent represented potential connectivity between sites and may influence northern bristlebird persistence.

Because of their low dispersal capability (Baker, 2000) and the dynamic nature of their habitat, northern bristlebird presence may also be strongly influenced by landscape context, such as the availability of rainforest refugia or creek gullies that could provide short term refuge following habitat loss. To account for this, the distance from the centroid of each site to nearest rainforest margin was measured. A watercourse lines GIS dataset (obtained from Queensland Department of Natural Resources and Mines, 2016) was also used to measure the distance between sites and nearest creek line refuge.

3.3.5. Statistical analysis

We collected data on a wide range of vegetation characteristics. To address our hypotheses, we needed to identify a small, meaningful set of predictor variables. We first investigated the pattern of correlation between the variables using Pearson’s correlation tests. Predictor variables with a high correlation essentially describe the same pattern in the response, and therefore can be safely removed before null hypothesis testing (Zuur et al., 2010). For paired predictor variables with a correlation coefficient of >0.5, the variable with the least effect on northern bristlebird presence was excluded. All statistical analyses were conducted in R (version 3.3.2; R Development Core Team 2014).

Plant community composition

We tested whether there were any differences in plant community composition between historic and currently occupied sites. An ADONIS multivariate analysis was performed using the species cover estimates in the Vegan R package, as it is considered to be more robust than an analysis of similarity (ANOSIM) (Oksanen et al., 2008). Differences in plant community composition between northern bristlebird presence/absence and region were tested for statistical significance compared against 1,000 null permutations. Patterns in plant community composition were then graphically represented as a two-dimensional non-metric multi-dimensional scaling (NMDS) ordination plot (using Bray-Curtis distances; stress = 0.201).

Habitat attributes

Generalized linear mixed-effects models (GLMMs) were used to model the relationship between habitat predictors and northern bristlebird presence. These models provide a robust statistical method for nested study designs, where data are hierarchically structured (Zuur et al., 2010). All habitat variables were standardized to unit mean and standard deviation to improve model convergence. Models were constructed based on five main hypotheses for explaining northern bristlebird presence: (1) foraging structure, (2) nesting structure, (3) shrub and canopy structure, (4) refuge availability, and (5) habitat extent (Table 3.2).

Spatial autocorrelation in northern bristlebird presence was tested using a join count statistic to accommodate presence/absence data. The occupied sites were found to be spatially clustered in the study region (z = 15.105, p < 0.001). Generally, sites within 10 km2 had similar occupancy. The spatial pattern of decline in northern bristlebirds has resulted in currently occupied sites being highly clustered. Since our aim was to inform reintroduction efforts that require information on habitat attributes for persistence, we did not examine the relative importance of spatial autocorrelation (e.g. distance to nearest occupied patch). Instead, to ensure habitat effects were not masked by spatial autocorrelation, we generated a spatial auto-covariate and included this variable in all models (including null models) (Crase et al., 2012). To test whether the inclusion of the spatial auto-covariate variable adequately dealt with spatial autocorrelation in models, a correlogram was produced to examine remaining spatial autocorrelation of model residuals using Moran’s I (Appendix 3).

These models were fitted with a logit link using the lme4 R package (R Development Core Team, 2014). To accommodate the hierarchical structure of the data, a variable indicating the region that each site belonged to was included as a random effect. All models were fitted using maximum likelihood.

An information theoretic approach was used to assess the relative support for each of the five hypotheses. Specifically, the Akaike’s Information Criterion (AICC) statistic was calculated for each model. Corrected AIC (AICC) statistics were used to account for the small number of sites in our analysis. The models were then ranked according to the relative differences in these statistics (∆AICC) and Akaike weights (ωi) were calculated.

The relative performance of each model was compared to identify the model that was best supported by our data.

Table 3-2. Habitat patch variables used for predicting northern eastern bristlebird presence across their historic range.

Variable Code Description Hypothesis 1: Foraging structure Composite variable of Grass height (cm) x grass cover Grass volume GV (%) 90% grass quantile GH.q Value of the 90% grass height quantile (cm) NMDS score of plant composition found within the ground Grass composition grass.NMDS vegetation stratum (0-1m) Hypothesis 2: Nesting structure Mean cover of Sorghum leiocladum within the 0-1m Sorghum cover SOR vegetation stratum (%) Mean cover of Poa spp.within the 0-1m vegetation stratum Poa cover POA (%) Value of the 90% Sorghum leiocladum height quantile 90% Sorghum quantile SOR.q (cm) Hypothesis 3: Shading structure Total summed cover of all species cover estimates within Total shrub cover TS the shrub vegetation layer (1-3m). Could be >100% Total canopy cover TC Total estimated cover of the canopy layer (>3m) NMDS score of plant composition found within the shrub Shrub composition shrub.NMDS vegetation stratum (1-3m) Hypothesis 4: Refuge availabilty Distance to the nearest rainforest margin (2013 aerial Distance to rainforest DRF imagery) Distance to creek/gully DC Distance to the nearest creek/wet gulley

Hypothesis 5: Habitat size Extent of habitat with grass dominated understorey (ha). Grassy habitat size GHS See Table 2 for full description Extent of habitat with shrub dominated understorey (ha). Shrubby habitat size SHS See Table 2 for full description

Classification tree analysis

Classification tree analysis is a flexible, robust analytical method that is capable of dealing with nonlinear relationships, high-order interactions and missing values (De'ath & Fabricius, 2000). Additionally, classification trees are robust against multicollinearity, and have greater ability to detect factor interactions than traditional GLM models (Vayssières et al., 2000). Classification trees therefore provide a good complementary analysis to

standard habitat models when a complex set of predictors is involved (Andersen et al., 2000).

A full classification model was fitted using all vegetation and spatial predictor variables using the rpart R package (Milborrow, 2016). Once the full model was fitted, trees were pruned using 10-fold cross validation (Breiman et al., 1984). Model performance was evaluated by area under the receiver operating characteristic curve (AUC) value (Fielding & Bell, 1997).

3.4. Results

3.4.1. Plant community composition

The ADONIS showed that plant community composition differed significantly between occupied and unoccupied sites (p=0.003), but plant community composition explained a relatively small proportion of the variation in presence (r2=0.219). Region had no significant effect on northern bristlebird presence (r2=0.162, p=0.273). NMDS ordination shows that there was no pattern in vegetation community across occupied and historically occupied bristlebird sites (i.e. bristlebird presence) or by region group (Fig. 3.2).

3.4.1. Habitat attribute models

Preliminary analysis showed northern bristlebird sites had significant spatial clustering in the landscape, meaning sites that were closer displayed similar patterns in habitat attributes and northern bristlebird occupancy. A correlogram showed that without a spatial autocovariate function, model residuals were significantly spatially autocorrelated (Appendix 3). With the spatial autocovariate added, spatial autocorrelation of model residuals was no longer significant.

Based on Akaike weights (ω), habitat extent was the best model for explaining northern bristlebird presence (Table 3.3). Habitat extent was the only model to outperform the null (∆AICc = 10.477, ω = 0.981). Larger habitat patches were more likely to contain sites in which northern bristlebirds were still present, once spatial autocorrelation was taken into account.

Figure 3.2. Two-dimensional ordination plot of plant community composition of currently (black) and historically occupied (grey) northern eastern bristlebird sites classified by region.

Table 3-3. Relative performance of six models, representing five alternative hypotheses about the drivers of northern eastern bristlebird persistence across their historical range.

Hypothesis Rank Model Variables d.f. logLik AICc ∆AICc ω Habitat extent 1 m6 Grassy habitat extent 4 -2.502 13.263 0 0.981 Grass Grass volume, Grass height 2 m2 6 -5.353 23.259 9.996 0.007 Structure 90% quantile, grass community null 3 m1 nil 3 -8.793 23.74 10.477 0.005 Nesting Sorghum cover, Sorghum 4 m3 6 -5.963 24.478 11.215 0.004 Structure height 90% quantile, Poa cover Shrub Total shrub cover, Total canopy 5 m4 6 -6.514 25.581 12.318 0.002 Structure cover, Shrub community Distance to rainforest margin, Refuge 6 m5 5 -7.971 26.333 13.07 0.001 Distance to creek

3.4.2. Classification tree

Of the 49 potential predictor variables included in the classification tree analysis, only four variables were selected by the algorithm (after pruning to reduce overfitting) to include in a final classification tree. Crofton Weed (Ageratina adenophora) cover occurred twice (Fig 3.3). This five split tree performed well against the data with an AUC value of 0.87.

The probability of northern bristlebirds occurring at a site was greatest when Crofton Weed cover was >0.31% (essentially, when Crofton Weed was present), mean grass height was <44 cm, and total summed cover of all grass species was >141%. All sites with this combination of attributes had northern bristlebirds present. The second best variable set for predicting northern bristlebird presence only involved two variables: Crofton Weed cover >0.31% and mean grass height >44cm, in which case summed grass cover was not important. Sites with poorer grass quality (low grass height and cover) were predicted to have northern bristlebirds present only if Crofton weed did not exceed 13% cover.

36% 100 TRUE (T) Crofton weed FALSE (F) cover <0.31 (%) 57% 60 T F Grass height 44% <44 (cm) 43 T Total grass F cover <141 (%) 36% 37 T Grassy habitat F extent ≥ 48 (ha) 61% 11 T Crofton weed F cover ≥ 13 (%) 16% 24% 0% 79% 100% 89% 40 26 3 9 6 17 Figure 3.3. Classification tree showing habitat attributes that explain northern eastern bristlebird presence. The tree is read by following the left split if statement is true and following the right split if statement is false. Each node box shows the probability of northern bristlebird presence. For example, 40% of our data (15 sites) had <0.31% Crofton weed cover (1st left split), and a 16% probability of northern bristlebird being present.

3.5. Discussion

Northern bristlebird persistence was most strongly influenced by the extent of continuous grassy habitat within which the site was located. However, the classification tree analysis revealed that this relationship was contingent on the particular grass structural attributes of a site. Birds were more likely to be present at sites with higher and denser grass, but if there was no tall, thick grass cover, then larger habitat patches were actually less likely to support their persistence. Although currently occupied and historically occupied northern bristlebird sites did differ significantly in plant community composition, this only accounted for a small amount of the variation. Persistence of northern bristlebirds therefore appears to be highly dependent on large areas of grassy habitat with thick or high grasses.

The extent of contiguous grassy habitat was overwhelmingly important in predicting northern bristlebird presence. Northern bristlebird presence at sites with a larger extent of

grassy habitat may be a result of regional dynamics associated with fragmented metapopulations. In fragmented metapopulations, larger habitat patches are able to host larger populations with decreased risk of stochastic extinction (Hanski, 1999b). Northern bristlebirds are territorial (Gubler et al., 2016), so large habitat patches may be necessary to support a viable, self-sustaining populations. Population persistence will also depend on dispersal between patches, with stochastic extinctions likely even if habitat quality is high (Hodgson et al., 2009). Smaller patches are less likely to be located and recolonised by dispersing birds (Andren, 1994; Hanski, 1994).

The large changes in habitat quality that occur in highly dynamic ecosystems can exacerbate stochastic extinctions that are associated with patch isolation (Hanski, 1999a). The grassy patches of sclerophyll forests in this region depend on disturbance, with a 3-6 year fire interval currently recommended to maintain a thick grassy understorey in this highly productive region in which vegetation growth rates are fast (Tasker et al., 2016). Immediately following fire events however, cover in grassy patches is greatly reduced, and northern bristlebirds may need to disperse across an uninhabitable rainforest matrix to other grassy patches. If suitable patches are too isolated from each other, local extinctions may occur during the time a patch needs to regenerate following fire. For the northern bristlebird, reduced ability to disperse between small isolated patches has likely contributed to their highly fragmented population, and as a consequence the potential for stochastic extinctions is high.

Small habitat patches may also be more vulnerable to the effect of a high intensity wildfire than larger patches. A single high intensity fire can have a lasting effect on habitat structure (Schoennagel et al., 2008; Williams & Bradstock, 2008). The heterogeneity and lower fuel loads associated with a large patch of grassy forest may reduce the severity of a wildfire and leave a mosaic of different aged remnant areas for refuge within a single habitat patch (Parr & Andersen, 2006). This may also be important for the persistence of a dispersal limited species such as the eastern bristlebird.

In addition, the heterogeneity found within larger habitat patches may provide larger gradients of microclimate and food resource availability (Karr & Freemark, 1983). As a result, individuals may have access to more beneficial microhabitats, which in turn could

have more optimal or consistent thermal/moisture conditions for invertebrate food sources during times of environmental stress (Karr & Freemark, 1983). A higher abundance of microhabitats within large habitat patches may thus allow for greater reliability of food resources across multiple spatio-temporal scales. In addition, plant phenology is known to vary significantly with altitude and local microclimates, therefore larger habitat patches with a wider range of topographic features may have more variation in flowering and seeding (Rathcke & Lacey, 1985).

While the size of the grassy habitat patches within which northern bristlebird sites were located was important, classification tree analysis suggested this was contingent on the condition of the grass layer. Sites with taller and denser grass cover were more likely to still support northern bristlebirds. If sites had shorter, less-dense grass cover, then larger patch size was not a predictor of northern bristlebird persistence. Persistence of northern bristlebird within habitat patches therefore seems to be dependent on the maintenance of large areas of tall, dense grass.

For most birds, the presence of suitable nesting sites is a major determinant of habitat suitability (Lamb et al., 1993). Northern bristlebird nests are predominantly constructed in or near the base of large, tall tussocks generally comprised of Sorghum leiocladum or Poa spp., with occasional nests also found in Lomandra spp. All these species have thick clumping structures; mature S. leiocladum clumps can be 60-150 cm in height (Lazarides et al., 1991). Previous monitoring of breeding territories and reports from captive populations reveal nests are constructed 10-45 cm off the ground (Gubler et al., 2016; Holmes, 1989). This nesting behaviour is therefore consistent with our finding that northern bristlebird presence is associated with taller grass.

Ground cover is important foraging habitat for many ground-foraging, insectivorous birds, and is particularly critical during nesting (Antos et al., 2008). In captivity, an increase in foraging of live invertebrates has been observed by nesting individuals (Gubler et al., 2016). Nesting is an energetically demanding activity, and therefore abundant food resources are necessary for successful nesting and chick rearing. Antos and Bennett

(2006) found little plasticity in the foraging actions of ground-foraging woodland birds, suggesting consistency in invertebrate resources may be an important determinant for foraging. Invertebrate communities are known to fluctuate significantly with moisture levels (Frith & Frith, 1990), and high rainfall events have been shown to stimulate northern bristlebird breeding (Gubler et al., 2016). With breeding observed across seven months in the wild, the northern bristlebird breeding season is relatively long (from as early as August until February) and a high abundance of food resources is likely required for the duration (Gubler et al., 2016). Thick, tall grass cover may provide more consistent soil moisture conditions for invertebrates (Newell, 1997), allowing them to occur in adequate abundance throughout the breeding season.

All habitat patches with high probability of northern bristlebird presence contained crofton weed. Despite this, it is unlikely that Crofton weed is a critical factor for their persistence, and may reflect - instead - other factors with which both are associated. Both the spread of crofton weed and the health of many native grasses are associated with disturbance. In the recent past this would have been farmer burning associated with cattle grazing and logging. In earlier times it was probably indigenous burning practices. Many native grasses benefit from frequent fire, which kills competing shrubs, returns nutrients to the soil, removes dead matter and promotes resprouting (Certini, 2005; Fisher et al., 2009; Higgins et al., 2000; Morgan & Lunt, 1999). It is likely that this also benefits Crofton weed and is responsible for its prevalence in disturbed agricultural systems (Auld, 1970). Previous studies have note northern bristlebird foraging within dense areas of Crofton Weed (D. Charley, pers comm).

Overall, the extent of contiguous grassy habitat appears to be a key component for the long-term persistence of the northern bristlebird, contingent on quality of the grass layer present. Our results indicate that future reintroductions of this population should be into larger habitat patches, close to occupied territories, where a thick, tall grass structure is present. Maintenance of these sites using appropriate fire management should also be prioritised: frequent fire in highly productive eucalypt forests has been shown to promote an open grassy understorey (Tasker & Bradstock, 2006). By focussing management actions within these sites, particularly appropriate fire regimes, we may be able to

supplement and aid natural recovery of the wild population in habitat that is capable of supporting them over the longer term.

3.6. Acknowledgements

ZLS was supported by an Australian Government Research Training Program Scholarship (RTP) scholarship, the Eastern Bristlebird Recovery Northern Working Group and top up funding from the NSW Office of Environment and Heritage. This project, and a broader study of which it was a part, was carried out under the auspices of the NSW Environmental Trust project 2012/RD/21 and we gratefully acknowledge their support. Addition funding has been provided by the Saving our Species program of the NSW Office of Environment and Heritage; Birds Queensland; Birdlife Australia; and the University of Queensland. We would like to thank NSW OEH Ecosystems and Threatened Species North East, QLD and NSW National Parks and Wildlife Services for their advice and assistance, Dr Simon Blomberg and Jeffrey Hanson for statistical advice, and the landowners on whose land most of the remaining birds are found for their support of our project and the eastern bristlebird.

Chapter 4

Patterns of invertebrate food availability and the persistence of an insectivore on the brink

Submitted to Austral Ecology

Plate 3. Invertebrate pitfall trap beneath dense Sorghum tussock at Spicers Gap, Main Range.

4.1. Abstract

Globally, insectivorous birds are at high risk of decline. One explanation of this relates to changes in invertebrate resources due to anthropogenic pressures. The northern population of the eastern bristlebird relies heavily on invertebrate food resources, and has experienced an 80% population reduction over the past 40 years. We investigated invertebrate abundance and nutritional quality across 23 currently and historically occupied northern bristlebird sites to determine whether extant territories were associated with more, or more nutritious, invertebrate resources. Pitfall and leaf-litter invertebrate sampling was done in both breeding and non-breeding seasons from 2014 - 2016. There was no difference in abundance, biomass or nutritional value of invertebrates between occupied and abandoned territories, however, within territories invertebrate abundance and nutritional value did correspond to the habitat characteristics with which bristlebirds are associated. Nutritional value of invertebrates increased with proximity to rainforest, while the abundance of macro-invertebrates (>1 mm) was correlated with grass height. Bristlebird territories are often close to rainforest margins, and these ecotones may provide more nutritious mesic-associated invertebrates. Higher abundances of large invertebrates in tall grasses may also contribute to the known association of bristlebirds with tall grasses. Maintenance of tall grass adjacent to rainforest through appropriate fire and grazing management is likely to be important for northern bristlebird recovery and long- term persistence of the population.

4.2. Introduction

Insectivorous birds are sensitive to land use change, habitat loss, degradation and fragmentation (Bennett & Watson, 2011; Watson, 2015). This is particularly true for ground-feeding insectivorous birds, which are experiencing some of the greatest declines globally (Antos & Bennett, 2006; Benton et al., 2002). Anthropogenic disturbances within grassy native vegetation, such as land use change, habitat fragmentation, introduction of exotic species and changes to natural fire regimes can have a strong influence on local habitat conditions and subsequently the invertebrate communities they are able to support (Andersen, 1988, 1991; Chambers & Samways, 1998; Didham et al., 1996; Pryke & Samways, 2012; Swengel, 2001; Zanette et al., 2000). Some research has suggested that

changes in prey availability or accessibility within native ecosystems may be a leading contributor to insectivore decline (Razeng & Watson, 2015; Rioux Paquette et al., 2014; Watson, 2011).

The eastern bristlebird (Dasyornis brachypterus) is an endemic small, cryptic, cover dependent passerine of Southeast Australia with limited flight ability (Hartley & Kikkawa, 1994). The northern population (henceforth referred to in this paper simply as the ‘northern bristlebird’), is critically endangered and isolated (> 800km) from the southern populations, occurring only along the border of New South Wales and Queensland in eastern Australia (OEH, 2012) (Fig 1b). The northern bristlebird occupies patches of grassy sclerophyll forest on basaltic ridges and slopes, typically embedded within rainforest, and (now) often bordered by cleared land (Holmes, 1989; Lamb et al., 1993; Rohweder, 2000). This contrasts markedly with the southern populations of the eastern bristlebird, which occupy coastal heath (Baker, 1997, 1998b, 2009; Baker et al., 2012).

The northern bristlebird has declined by an estimated 80% in the last forty years (Garnett & Crowley 2000). The decline of the species in general (i.e. both northern and southern populations) has been attributed to habitat loss and fragmentation, introduced predators and inappropriate fire regimes (Bain et al., 2008; Baker, 1997; OEH, 2012). The 2016 annual census (the most recent undertaken) of the northern bristlebird estimated the current wild population at 40 individuals, with just five confirmed breeding pairs (Wildsearch Environmental Services, 2016b).

Although eastern bristlebirds are considered omnivorous, consuming a range of invertebrates and seeds (Gibson & Baker, 2004; Gubler et al., 2016), faecal analysis and observations of foraging birds indicate their diet is predominately invertebrates (Gibson & Baker, 2004). The major invertebrate groups consumed in the southern population studied by Gibson & Baker were Coleoptera (beetles), Hymenoptera (ants), Diptera (flies) and Mantodea (mantids), with a lower incidence of Arachnids (spiders). Anecdotal observations of northern birds in the field and in captivity indicate their diet is similar.

Omnivorous bird species often increase their dependence on insect prey during the breeding season as insects provide important nutrients necessary for successful breeding (Razeng & Watson, 2015). As a result, breeding success and recruitment can be influenced by availability of invertebrates (Bryant, 1975; Eeva et al., 1997; Martin, 1987; Robb et al., 2008). Knowing whether the decline in the northern bristlebird is at least partly a response to a decline in food availability in grassy sclerophyll forest patches would aid managers in determining management actions to maintain high quality habitat for the northern bristlebird. This was the rationale for this study, in which we compare invertebrate resources between current and historically occupied northern bristlebird sites, and between the microhabitats found within habitat patches, in an attempt to determine whether disappearance of northern bristlebirds may be in part a response to fewer invertebrate resources.

Specifically, we first tested whether northern bristlebird presence in patches was correlated with invertebrates. If northern bristlebird disappearance from a site is associated with insufficient food, we might expect historically occupied (i.e. abandoned) sites to have lower abundance, biomass or nutritional value of invertebrates, or different invertebrate community composition, compared to currently occupied sites.

We also tested whether invertebrate patterns within habitat patches were related to specific vegetation structural attributes or patch characteristics. Northern bristlebird presence has been shown to be associated with larger densely grassy patches (Stone et al., 2018). If this association is a response to prey availability, then we might expect patches with taller, thicker grass structure to have more prey. We also tested whether invertebrate abundance or composition differed between breeding (spring) and non- breeding (autumn) times of year, to see whether currently occupied sites had a higher, or more nutritious, invertebrate profile during the breeding season than historically occupied sites do.

4.3. Methods

3.3.1. Study area

Fieldwork was carried out at all known territories where grassy habitat still exists and is currently, or has historically been, occupied by northern bristlebirds. Bristlebird sites are generally isolated from one another by denser forest, with regional clusters of sites found in and around Conondale, Main Range, and Lamington national parks in Queensland and Border Ranges National Park in NSW (Figure 4.1). The climate in the study area is subtropical-temperate, with warm, wet summers and mild, dry winters (Stern et al., 2000). Rainfall in the Border Ranges averages 947 mm/year, most of which falls during the summer, and air temperatures average 32 °C in summer and 22-26 °C in winter (Bureau of Meteorology, 2014). Sites are found clustered within small areas in and adjacent to the above-mentioned parks, largely on upper slopes or exposed ridgelines where rainforest gives way to open, grassy eucalypt forest or woodland. Much of the grassy eucalypt woodland found on lower slopes in the study region has been extensively cleared for agriculture, while rainforest is the dominant vegetation at higher altitudes.

3.3.2. Northern bristlebird presence and absence

Northern bristlebird presence and absence was determined using data from the biennial monitoring by Office of Environment & Heritage (Sandpiper Ecological Surveys, 2012; Wildsearch Environmental Services, 2015b, 2015c, 2015d, 2016b). These surveys consist of an initial five minute listening period followed by a five minute period broadcasting northern bristlebird recordings, and then a 20-25 minute listening period to pick up responses. Once birds respond, the survey ceases. Sites in which birds have not been detected for five years are classified as unoccupied (D. Charley pers. comm.). Since 2013 this standard monitoring has been supplemented by use of a specialised detector dog at the NSW sites. The dog has been shown to be able to detect as little as a single bristlebird feather or the scent of a bird even if the bird is not present at the time. This detector dog confirmed that the absence of birds at historically occupied sites recorded by the annual monitoring are indeed true absences (and not just a lack of detection or lack of calling). The combination of monitoring techniques allows us to have a high degree of confidence in the presence/absence records, despite the cryptic nature of the northern bristlebird.

153 °E

a Conondale Range b

NT QLD WA 27 °S SA NSW

VIC Brisbane TAS

Bristlebird territory Site 1 c Survey plot

28 °S Main Range Lamington Mt Barney

Border Ranges Site 2

3.3.3. Invertebrates

Invertebrate samples were collected at 23 of the 38 known current and historic northern bristlebird sites. While vegetation assessments were undertaken at all 38 sites, access constraints and planned burns meant that some sites were inaccessible during some invertebrate sampling periods. As a consequence, we have included only the 23 sites that were available for all sampling periods. Of the 23 sites, 9 sites are currently occupied, while the remaining 14 are sites in which northern bristlebirds used to occur but have not been detected for many years, and are now considered extinct. The number of plots surveyed at each site varied depending on site size. Northern bristlebird territories can be tightly concentrated to specific areas within a habitat patch or be spread over a larger area

of the habitat patch (i.e. Fig 1c). Larger sites had more survey plots: there were 4-6 plots at each site depending on size. To account for variation in the number of plots, trapping effort (‘trap nights’ = number of traps x number of nights) per site was calculated and included in models.

Sites were defined using the territory mapping of Holmes (1982), with the extent of home ranges taken from the annual OEH monitoring. A random sampling approach was used to locate invertebrate survey plots within sites, with each site defined in ArcGIS (ESRI, 2015). Each plot covered an area of 20 x 20 m and was randomly placed within 200 m of a known northern bristlebird location at a site. Each plot was further subdivided into sixteen 2.5 x 2.5 m subplots, of which two were randomly chosen for pitfall and leaf litter sampling for any given sampling occasion (Fig. 1c).

Observations of eastern bristlebird foraging have shown birds peck at and turn leaf litter with little to no digging below the surface (Gibson & Baker, 2004). To reflect this, invertebrates were collected using methods that target epigaeic (surface active) invertebrates: pitfall traps and leaf litter sampling (Sabu & Shiju, 2010). Sampling was conducted in 2015 and 2016 in May-June and September-October. These times were chosen to sample the non-breeding winter period (May-June) and the onset of breeding activity (September-October; Gubler et al., 2016).

Pitfall trapping

For each sampling period, one pitfall trap was centrally placed within a single randomly predetermined subplot (i.e. there was one pitfall trap per plot, giving a total of 4-6 per site per sampling period). Because of the presence of northern bristlebirds and other small threatened vertebrate species, wire bird netting with a 1.3 cm hex aperture was placed over each trap to avoid non-target bycatch. While this may have had some influence on the abundance of larger macro-invertebrates captured, the invertebrates eaten by the eastern bristlebird are considered to be mostly smaller than this (Gibson & Baker 2004) and this method was deemed necessary to avoid vertebrate mortality.

Each trap consisted of two round straight-sided plastic containers approximately 8 cm in diameter and 12 cm deep, one placed inside the other, the interior one with a removable lid. The use of nested containers allowed for removal of samples without disturbing surrounding soil. Traps were dug in so the rim was flush with the surrounding soil. Containers were left sealed for one week prior to sampling to reduce digging-in effects (Greenslade, 1973). Pitfall traps were then opened and filled with approximately 3 cm of a diluted propylene glycol fluid (1:3 parts water) and detergent to remove surface tension (Jud & Schmidt-Entling, 2008). A cover approximately 18 cm in diameter was positioned 10 cm above each trap to protect it from rain and flooding. Traps were then left open for seven consecutive nights after which they were removed and sealed. The contents of the traps were transferred to 70% ethanol to preserve them until sorting.

Leaf litter samples

At each sampling occasion a single leaf litter sample was also taken in each plot at a randomly predetermined subplot (excluding the pitfall trap subplot). This meant there was a minimum of 5 m between the pitfall trap and leaf litter sample, to avoid interactions (Ward et al., 2001a). Each litter sample was obtained by gathering leaves, litter and loose humus from a 50 cm2 quadrat dropped in the centre of the target subplot. Samples were collected quickly and carefully and stored in fine weave cotton bags to prevent escape of invertebrates. Leaf litter samples thus sampled the upper organic litter layer plus the loose humus layer that bristlebirds forage in. No underlying compact soil was included. Litter samples were passed through 5 cm wire sieves to separate larger litter material (i.e. sticks) and remaining material was emptied into a Berlese extraction apparatus (16.5 cm diameter, 18 cm height, 1.3 cm mesh, 25w filament lamp) and left for one week, as described by Sabu and Shiju (2010). Large litter material was checked visually for macro- invertebrates that may have been missed. Collection jars were attached to the base of the Berlese apparatuses and filled with 20 ml of 70% ethanol solution.

Invertebrate identification

Invertebrates captured from each site were sorted to order level using a dissection microscope (Leica EZ4HD) and the dichotomous key of Harvey and Yen (1997). Taxa were grouped and identified following the protocols of Razeng and Watson (2012). Because the diet of the northern bristlebird has been little studied but the available observations show they consume very small insects (A. Beutel pers. comm.), we counted invertebrates in both macro- (>1 mm) and micro- (<1 mm) size classes. We counted the total number of individuals (“total abundance”) in each order per size class for each sample. Samples were then dried in a 60 ° C oven for 48 hrs to obtain the dry weight (biomass; accurate to the nearest 0.0001 g).

3.3.4. Habitat patch characteristics

To test whether vegetation structure influenced invertebrate abundance or composition, we measured grass height along two 20 m transects that crossed the centre of each survey plot (Fig 1c). Cover (to the nearest 5 %) was estimated for each species in the ground (0-1 m), shrub (1-3 m) and canopy (>3 m) vegetation layers for each 10 x 10 m survey plot quarter. These measurements provided estimates of individual species cover and were also used to calculate overall ground, mid-story and canopy cover. The extent of grassy habitat at each site (i.e. the extent of suitable grassy habitat over the contiguous habitat patch in which a site is located) was mapped by visually delineating grassy patches using ADS40 aerial imagery (taken in 2009, the most recent available). This imagery has a resolution of 5 m. The distance from the centre of each survey plot to the nearest rainforest margin was also measured from the aerial imagery.

3.3.5. Statistical analysis

Invertebrate and habitat data were averaged across plots for each site. Invertebrate community composition was summarised in non-metric multidimensional scaling (NMDS) scores, using the total abundances from each combination of order and size class (micro- and macro-invertebrates). Total abundance and biomass variables were calculated from the sum of both pitfall and leaf litter samples. The nutritional value of the invertebrates at each site was calculated using the total abundance of macro-invertebrates from five main

invertebrate groups, Araneae, Coleoptera, Diptera, Hemiptera and Hymenoptera, multiplied by the crude protein and fat values (mg/individual) from Razeng and Watson (2015). Their study analysed the macro- and micro-nutrients for the eleven main invertebrate groups in open woodland forests in South Australia. While this is not the same as the habitat in this study, it is the closest available analysis of invertebrate nutritional composition, and represents the potential nutritional value available within a grassy, eucalypt-dominated habitat type. We included only the five mentioned invertebrate groups in our calculations because: they are known eastern bristlebird prey (Gibson & Baker, 2004), they are known to be preferred prey for insectivorous birds more generally (Moorman et al., 2007; Razeng & Watson, 2015), and they were abundant enough to be present in most of our samples (comprising 58 % of the sampled macro-invertebrates).

We used an information-theoretic approach to assess which aspects of invertebrate resources bristlebird presence was correlated with. Generalized linear mixed-effects models (GLMMs) were fitted with a logit link using the lme4 R package (R Development Core Team, 2014). All invertebrate and habitat predictors were standardised before including in models. We controlled for spatial auto-correlation of northern bristlebird sites by incorporating an auto-covariate into each candidate model, including the null models (Crase et al., 2012). Variation in sampling effort between sites was controlled by adding the log of trap effort as an offset in all models.

To investigate whether there were differences in invertebrates between current and historically occupied sites at different times of year, we tested whether there were any differences in invertebrate abundance, biomass or nutritional value between spring and autumn for current and historically occupied sites using paired t-tests. We then tested for differences in invertebrate community composition between either sites or seasons using an ADONIS multivariate analysis in the Vegan R package (Oksanen et al., 2008), and tested for statistical significance by comparison against 1,000 null permutations.

Finally, we tested whether invertebrate characteristics at sites were related to habitat attributes (not just a simple ‘occupied’ versus ‘historically occupied’ patch comparison).

Candidate models were chosen a priori to reflect four main hypotheses to explain bristlebird presence: (1) total invertebrate abundance, (2) total invertebrate biomass, (3) invertebrate community composition and (4) invertebrate nutritional value. For these models we fitted GLMMs with a Poisson error distribution for invertebrate abundance as a response variable against grass structure, mid-story and canopy cover and habitat patch context. We also looked at whether invertebrate biomass and nutritional value were influenced by these habitat predictors using a Gaussian error distribution with an identity link for biomass and nutritional value.

4.4. Results

A total of 41,413 arthropods was collected from 23 historically and currently occupied northern bristlebird sites (Table 4.1). Collembola, Acarina, Hymenoptera and Coleoptera made up 76 % of the invertebrates sampled. The number of invertebrates per site (per trap effort) ranged from 20 to 196 (77 ± 33 , n = 23) and total arthropod dry weight per site ranged from 0.20 g to 1.07 g (0.53 ± 0.20 g, n = 23) per sampling period.

3.4.1. Bristlebird presence in relation to invertebrate resources

Northern bristlebird presence and absence were modelled as a response to various invertebrate predictors (Table 4.2). Conditional and marginal R2 values were > 0.517. However, within our candidate model set, while nutritional value was the best performing model, it had only slightly better predictive power than the null model (∆AICc = 0.460).

Northern bristlebird presence also was not related to invertebrate community composition found at a site (r2 = 0.06, p = 0.06). While omnivorous bird species can often rely more on invertebrate resources during breeding, we found no relationship between northern bristlebird presence and invertebrate composition, abundance or biomass when just the spring breeding period was considered (r2 = 0.05, p = 0.21).

Table 4-1. Number of invertebrates collected in pitfall traps and leaf litter samples within northern bristlebird habitat patches.

% of total > 1mm < 1 mm Taxonomic group Common name n captures (macro) (micro) Acarina Ticks 10162 24.83 276 9886 Araneae Spiders 1083 2.65 805 278 Amphipoda Amphipods 533 1.30 525 8 Blattodea Cockroaches 166 0.41 163 3 Coleoptera Beetles 4640 11.34 1862 2771 Collembola Springtails 10014 24.47 1963 8051 Dermaptera Earwigs 235 0.57 234 1 Diplopoda Millipedes 178 0.43 177 1 Diplura Two-pronged bristletails 45 0.11 29 16 Diptera Flies 1676 4.10 1145 531 Gastropoda Snails 38 0.09 30 8 Geophilomorpha Soil centipedes 99 0.24 99 0 Haplotaxida Earth worms 85 0.21 85 0 Hemiptera True bugs 793 1.94 327 466 Hymenoptera Ants, bees, wasps 6338 15.49 5322 1016 Isopoda Wood lice 1514 3.70 1192 322 Isoptera Termites 4 0.01 4 0 Leipidoptera Butterflies, moths 18 0.04 18 0 Lithobiomorpha Stone centipedes 46 0.11 46 0 Mantodea Praying mantids 5 0.01 4 1 Onychophora Velvet worms 8 0.02 8 0 Orthoptera Crickets, grasshoppers 69 0.17 68 1 Phthiraptera Lice 1 0.00 0 3 Pseudoscorpionida Psuedo-scorpions 462 1.13 45 417 Scolopendromorpha Tropical centipedes 24 0.06 24 0 Scorpionida Scopions 51 0.12 51 0 Scutigerida House centipedes 15 0.04 15 0 Siphonaptera Fleas 4 0.01 4 0 Thysanoptera Thrips 12 0.03 8 4 Unidentified larvae 2565 6.27 1660 905 Unidentified 44 0.11 15 29 TOTAL 40927 16204 24928 *Invertebrate orders based on Harvey & Yen (1997) classifications

Table 4-2. Relative performance of alternative models testing northern bristlebird presence in response to invertebrate predictors in currently occupied and historically occupied sites.

Model Intercept Total Biomass NMDS1 NMDS2 Nutrients d.f. logLik AICc ∆AICc ω Macro- -3.818 1.302 4 -9.777 29.8 0 0.331 nutrients Null -3.631 3 -11.307 29.9 0.101 0.314 Biomass -3.746 -1.220 4 -10.240 30.7 0.926 0.208 Abundance -3.625 -0.339 4 -11.086 32.4 2.619 0.089 Composition -3.571 0.634 0.730 5 -9.872 33.3 3.498 0.058

3.4.2. Invertebrate patterns within bristlebird habitat patches

We found no significant differences in the total composition (r2 = 0.002, df = 2, p = 0.75), spring composition (r2 = 0.05, df = 2, p = 0.21) or autumn composition (r2 = 0.02, df = 2, p = 0.42) of invertebrates across current and historically occupied sites (Fig 4.2). There was also no difference in the abundance, biomass or nutritional value of invertebrates across seasons or between current and historically occupied sites (Fig 4.3). Although average abundance, biomass and nutritional value were slightly higher during spring, these differences were not significant.

We found no relationship between northern bristlebird presence and invertebrate variables at the time of sampling, but invertebrates themselves were strongly influenced by habitat variables, particularly at the patch scale (Table 4.3). Patch context (distance to rainforest and patch size), grass structure and mid-story/canopy cover were all good predictors of nutritional value, with all models used to predict nutritional value of a northern bristlebird habitat patch outperforming the null model (Table 4.3). The model most strongly supported by the data included patch context, with sites closer to the rainforest margin having invertebrates with a higher nutritional value than sites further from the margins (Fig 4.4). This may be because ants have relatively low nutritional value because of their high chitin content (Redford & Dorea, 1984), and we found their abundance (and proportional representation) increased with distance from the rainforest margin and habitat patch size.

Model Rank Equation d.f. logLik AICc ∆AICc ω 1. Invertebrate abundance Grass structure 1 1.580 + (0.134 * GrassH) + (0.176 * GrassC) 4 -176.13 362.48 0 0.55 Canopy structure 2 1.595 + (-0.062 * Shrub) + (-0.201 * Canopy) 4 -176.33 362.87 0.399 0.45 Patch context 3 1.618 + (0.001 * Patch) + (0.158 * DistRF) 4 -189.77 389.77 27.29 0 Null 4 1.596 2 -206.69 417.97 55.5 0 2. Invertebrate biomass Null 1 -2.242 3 -17.415 42.093 0 0.92 Grass structure 2 = -2.223 + (0.197 * GrassH) + (-0.128 * GrassC) 5 -17.569 48.667 6.575 0.04 Canopy structure 3 = -2.256 + (-0.153 * Shrub) + (-0.126 * Canopy) 5 -17.574 48.677 6.584 0.03 Patch context 4 = -2.177 + (0.195 * Patch) + (-0.020 * DistRF) 5 -19.13 51.789 9.696 0.01 3. Invertebrate nutritional value Patch context 1 269.009 + (-38.480 * Patch) + (11.406 * DistRF) 5 -121.6 256.74 0 0.62 Canopy structure 2 269.009 + (-26.714 * Shrub) + (19.214 * Canopy) 5 -122.52 258.57 1.83 0.25 Grass structure 3 269.765 + (-6.448 * GrassH) + (-14.200 * GrassC) 5 -123.16 259.85 3.107 0.13 Null 4 269.009 3 -131.21 269.69 12.95 0

Patch size was not as important in predicting the nutritional value of sampled invertebrates, with only a weak negative relationship (i.e. larger patches had slightly lower nutritional value of invertebrates). Canopy also influenced invertebrate nutritional value, with sites with an open canopy and lower shrub cover having higher invertebrate nutritional value.

Grass structure had a strong influence on overall invertebrate abundance within northern bristlebird sites (Table 4.3). While grass height alone had no effect on overall invertebrate abundance, this appears to be the result of opposite responses by macro- and micro- invertebrates (Fig 4.5). Macro-invertebrates increased in abundance with grass height, while micro-invertebrates decreased.

Figure 4.5. Total abundance of (a) macro and (b) micro invertebrates captured in northern bristlebird sites in relation to grass height.

While invertebrate abundance increased with grass structural complexity, invertebrate biomass did not. None of the alternative models for predicting invertebrate biomass

outperformed the null model (Table 4.3). This is likely to be explained by the contrasting patterns of micro- and macro-invertebrate abundance.

4.5. Discussion

Food limitation, particularly of invertebrate resources, has been suggested as a key determinant of insectivore abundance and distribution (Razeng & Watson, 2015; Watson, 2011; Zanette et al., 2000). However, we found no evidence of differences in invertebrate availability between historically and currently occupied bristlebird sites. Invertebrate abundance, biomass, nutritional value and community composition were all similar between sites in which northern bristlebird persist and those from which they have been lost. This suggests that food limitation is not a key driver in birds abandoning habitat patches. Main prey items such as Hymenoptera and Coleoptera (Gibson & Baker, 2004) were relatively abundant across all sites, making up a quarter of all invertebrates, and almost half the macro-invertebrates present.

However, we did find that there were significant effects of site variables, particularly at the patch-scale, on invertebrate patterns. Nutritional value of invertebrates was higher at sites closer to the nearest rainforest margin, and abundance was greater in habitat patches with taller, thicker grass, both habitat features with which northern bristlebird persistence is associated (Holmes, 1989; Lamb et al., 1993; Rohweder, 1999; Stone et al., 2018).

We found no evidence of differences in invertebrates between non-breeding and breeding periods. Breeding is a highly energetically demanding activity and successful breeding is often determined by food availability (Martin, 1987). For omnivorous birds, invertebrates often provide critical nutrients during the breeding season, and become a main dietary component at this time (Wilson et al., 1999). Invertebrate resource availability can decline rapidly throughout the breeding season with the lowest abundances occurring during nesting and fledging periods (Martin, 1987). According to anecdotal evidence, northern bristlebirds form strong pair bonds during the breeding season, and males have a high energy investment at the commencement of breeding through territorial calling, nest building and feeding females throughout the courtship period (A. Beutel pers. comm.).

Monitoring of invertebrate resources near known nest sites during the nesting period (not just within the territory more generally), compared to non-breeding locations and historic sites, may be useful to definitively rule out food limitation during breeding.

Despite finding no differences in invertebrate resources between current and historically occupied sites, northern bristlebird occupy areas with specific habitat characteristics, in particular those that are close to rainforest margins (within 400m) and where there is tall thick grass cover and a relatively open canopy (Lamb et al., 1993; Rohweder, 2000). We found higher invertebrate nutritional value and abundance of macroinvertebrates in these areas.

Stone et al. (2018) found a strong influence of habitat patch size on northern bristlebird presence, with northern bristlebirds more likely to persist in large patches of contiguous grassy habitat. While we found no effect of patch size on invertebrate resources per unit area, by simple extrapolation larger habitat patches should provide more invertebrate resources (since bird territories don’t extend into the rainforest). Not only do large patches provide a greater area for foraging, they probably also have greater habitat heterogeneity (Karr & Freemark, 1983). Habitat heterogeneity allows for a wider diversity of localised environmental conditions and produces more variation in spatio-temporal distribution of prey (Law & Dickman, 1998). This may contribute to a higher overall year-round availability of invertebrate resources in large habitat patches (Ford et al., 2001). The mountainous sub-tropical region in which the northern bristlebird is found has substantial seasonal and year to year climatic variation: during drought conditions summers may be very hot and dry, while more typical summers are warm and can have high rainfall due to the subtropical nature of the region. Having a greater range of microhabitats available within a patch and close to it may mean that large patches less often experience a food shortage than small patches. The observation that some birds have temporarily moved down into gullies during extremely hot dry summers (Charley and Baker, pers. comm.) is consistent with the possibility of movements tracking food resources.

Despite no relationship between northern bristlebird presence and nutritional value of invertebrates, sites closer to rainforest margins had a higher overall nutritional value of invertebrates than sites further from the margins. While our study was not designed to test the effects of ecotones on invertebrates, the almost universal proximity of northern bristlebird sites to rainforest margins suggests the ecotone is important. Previous explanations have suggested rainforest may provide an important short term refuge following fire events (Lamb et al., 1993; Rohweder, 2000). Our results suggest the ecotone may also affect invertebrate resources, potentially from a spill-over effect (Lacasella et al., 2015; Rand et al., 2006).

Grassy habitats are often subject to strong fluctuations in temperature and moisture and higher anthropogenic disturbance (Matlack, 1993). Invertebrates are highly sensitive to environmental conditions (Kremen et al., 1993): vegetation composition and structure, microclimate, physical and chemical soil properties, topography and disturbance occurring within habitat patches all play a role in shaping the invertebrate communities found within them (Curry, 1994). Higher soil nutrients, soil moisture and leaf litter along forest-grassland boundaries (Turton & Sexton, 1996) and diverse vegetation structure (Berg & Pärt, 1994) may contribute to higher abundances and quality of invertebrates. Lacasella et al. (2015) found that forest dwelling arthropod species were less sensitive than grassland species to edge effects, with many forest species found ‘spilling over’ into the ecotone area. By remaining close to rainforest margins, northern bristlebirds may be able to utilise rainforest for short term refuge and prey resources during dry periods. Few studies have looked at invertebrate patterns across natural rainforest-grassy forest ecotones, with most focussing on disturbed agricultural or closed forest-grassland edges (Durães et al., 2005; Kotze & Samways, 2001). The few studies that have been carried out at natural forest-grassland edges have found them to have higher numbers of predatory invertebrates (Ingham & Samways, 1996), carabid beetles and amphipods (Heliölä et al., 2001; Kotze & Samways, 2001; Yu et al., 2007). Differences in invertebrate resources between the rainforest and grassy forest may provide complementary nutritional resources which can be consumed at different periods depending on energetic needs (Law & Dickman, 1998).

Our data suggested that there was no significant difference in invertebrate community composition between sites or sampling periods. This may be as a result of only identifying specimens to order level. This level of identification was based on the generalist feeding patterns of bristlebirds observed by Gibson & Baker (2004), however there can often be significant within-order variation in invertebrate responses (Pik et al., 1999; Timms et al., 2013) even if no pattern is shown at a higher taxonomic level. Within order, there can be variation in the composition of families, genera and species in relation to, for example, habitat type, environmental conditions and disturbance (Andersen, 1991; Andersen & MÜller, 2000; Orgeas & Andersen, 2001; Radford & Andersen, 2012).

Tall, thick grass has been previously identified as a critical component of northern bristlebird habitat (Holmes, 1989; Lamb et al., 1993; OEH, 2012; Rohweder, 2000; Stone et al., 2018). In this paper, we found more macro-invertebrates in grass with greater height and cover. Plant density, vegetation diversity and patch attributes can significantly affect patterns of invertebrates (Denno & Roderick 1991). Greater macro-invertebrate abundance in taller grass may reflect this, and may be partially contributing to the importance of tall grass for northern bristlebirds. Larger invertebrates often provide higher nutritional value for foraging effort (Holmes & Schultz, 1988; Kaspari & Joern, 1993; McCarty & Winkler, 1999), and could thus allow less foraging effort, likely to be particularly important during nesting (which occurs in tall grass tussock patches).

Whilst Gibson & Baker (2004) found that seeds made up only a small proportion of bristlebird diet, their study was on the southern population which lives in heathland. It is possible that seed resources may be more important to northern birds. In grassy forest habitat, seeds may be more important to bristlebirds. Captive birds are provided with seed but is rarely consumed (A. Beutel pers. comm.). Future research on the actual diet of the northern bristlebird would help confirm the relative importance of seed and invertebrates at different times of the year, though the cryptic, ground-dwelling nature of bristlebirds, as well as their rarity, make this difficult. Observations of captive birds is possible, however captive raised individuals may not reflect wild dietary preferences or foraging behaviours (Håkansson & Jensen, 2005; Snyder et al., 1996). An indirect way to accomplish this could

be to compare seed availability between occupied and abandoned sites to see whether there are differences.

The results of this study suggest that altered invertebrate assemblages are unlikely to be a major determinant of current northern bristlebird distribution. We found no direct evidence to suggest currently occupied northern bristlebird sites have greater abundance, biomass, nutritional value or diversity of invertebrate prey per unit area. However, distance to rainforest and understorey vegetation structure may play more significant roles for northern bristlebird persistence, and this may be partly through their effects on food availability. In contrast, continued loss of northern bristlebird habitat and deterioration in grass condition due to inappropriate fire are a higher priority for management (Stone et al., 2018).

4.6. Acknowledgements

ZLS is supported by an Australian Government Research Training Program Scholarship (RTP) scholarship, funding from the NSW Office of Environment and Heritage, and the Eastern Bristlebird Recovery Northern Working Group. This study was undertaken under the auspices of the NSW Environmental Trust project 2012/RD/21 and we gratefully acknowledge their support. Funding has been provided by the Saving our Species program of the NSW Office of Environment and Heritage; Birds Queensland; Birdlife Australia; and the University of Queensland. We would like to thank NSW OEH Ecosystems and Threatened Species North East, as well as the QLD and NSW National Parks and Wildlife services for their advice and assistance, Dr Anita Cosgrove for her laboratory assistance, Dr Simon Blomberg and Jeffrey Hanson for statistical advice, and private landowners for their support of our project and eastern bristlebird recovery.

Chapter 5

Frequent fires are needed for maintaining grassy forest habitat for the northern bristlebird

To be submitted to International Journal of Wildland Fire

Plate 4. Planned burn at an occupied northern bristlebird site in the Border Ranges.

5.1. Abstract

Fire can play an important role in maintaining grassy forest habitats in the presence of potential encroachment by woody plant species. Absence of fire in grassy systems has been connected to changes in vegetation composition and the location of the rainforest- open grassy forest boundary. The northern population of the endangered eastern bristlebird occupies open grassy sclerophyll forest habitat close to rainforest margins. Local conservation managers have observed that its grassy forest habitat has been increasingly invaded by rainforest plants and weeds. We tested the degree to which this encroachment correlated with changing fire regimes by collating 30 years of fire history data for the 38 current and historically occupied bristlebird sites, and examining the relationship between fire history, bristlebird persistence and habitat condition. We found that the extent of patches historically occupied by bristlebirds declined by over 50 % in the last 30 years due to encroachment, and that this decline was strongly associated with longer fire intervals and extended fire absences. A lack of fire was linked to changed understorey condition, with reduced tussock cover and decreased grass volume (both key habitat requirements of the northern bristlebird) associated with longer fire intervals. Reduced fire frequency in northern bristlebird habitat appears to be a major cause of their decline, by significantly reducing habitat extent and condition. Fire intervals longer than 8- 10 years lead to grassy forest attaining a fire resistant state. Active management through prescribed ecological burns is crucial for the persistence of the grassy forests on which northern bristlebirds and other threatened species in the region are dependent.

5.2. Introduction

Fire is an ecologically important disturbance for many grass-dominated communities around the globe (Bond & Keeley, 2005; Bowman et al., 2009). The persistence of certain kinds of grassy habitats depends on regular disturbance to reduce competitively dominant shrubs and trees, without which changes in vegetation structure and composition to a different vegetation type can occur (Fairfax et al., 2009b; Fensham & Fairfax, 2006; Moxham et al., 2016). In forested grassy systems, fire can promote grass growth, suppress exotic and competitive shrub species and favour increased diversity of grasses, interstitial herbs and forbs (Hoffmann, 1999; Peterson & Reich, 2008).

Rainforest and shrub expansion into grassy ecosystems has been observed particularly in high rainfall, humid regions, and has been linked to reduced fire in these systems (Crowley & Garnett, 1998; Gooden et al., 2009; Harrington & Sanderson, 1994; Tng et al., 2012). For instance, a lack of fire has contributed to the decline in the extent and condition of montane grasslands in the Bunya Mountains in south-east Queensland, Australia (Butler et al., 2014; Fairfax et al., 2009a; Fensham & Fairfax, 2006; Moravek et al., 2013), not far from our study region. Such change in the location of the forest-grassland boundary (or ‘ecotone’) and the contraction and disappearance of grassy habitat may be contributing to the local extinction of species that rely on this dynamic habitat. High-elevation grassy forests provide critical habitat for a range of threatened species in north-eastern New South Wales and south-eastern Queensland, such as the northern bettong (Bettongia tropica), Hastings River mouse (Pseudomys oralis) and the eastern bristlebird (Dasyornis brachypterus), and changes to fire regimes may be contributing to their decline (Abell et al., 2006; Johnson & McIlwee, 1997; Pyke & Read 2002; Sandpiper Ecological Surveys, 2000).

The eastern bristlebird is an endangered passerine occurring in south-east Australia. It is a small, cryptic bird with limited flying ability, spending the majority of its time on the ground foraging in the leaf litter (Hartley & Kikkawa, 1994). The species has been described as cover-dependent, and is strongly associated with habitat that contains a dense low understorey, occurring in a range of heath and shrubland vegetation types in the south (Baker, 2000) through to grassy forest in its northern range (Holmes, 1989; Lamb et al., 1993; Stone et al., 2018). Although once occurring in a near-continuous distribution in coastal and adjacent ranges of eastern Australia, it now consists of isolated southern populations in Victoria and southern NSW, and a northern population (henceforth referred to as the ‘northern’ bristlebird) found more than 700 km north, on the Queensland-New South Wales border (OEH, 2012). The southern populations are small (500-2000 birds) but relatively stable, and their coastal and montane heath habitats are largely protected in conservation reserves (Baker, 1997, 2000; Bramwell et al., 1992; Clarke & Bramwell, 1998). In contrast, the tiny northern population of approximately 40 wild birds is found only in patches of high-elevation grassy sclerophyll forest habitat (Rohweder, 2000; Wildsearch

Environmental Services, 2016b). These patches are typically bordered by rainforest and – now – cleared grazing land. Fire has been suggested to be a direct threat to the small remaining population (Holmes, 1989, 1992; Lamb et al., 1993), but others have suggested that it may also be needed to maintain the grassy habitat (Rohweder, 2000, 2006; Sandpiper Ecological Surveys, 2000). There is likely an optimal fire interval that can maintain habitat and prevent loss of individuals from frequent, or high intensity fires (Garnett et al. 2011).

Recent studies show that grassy sclerophyll forest patches within the northern bristlebird’s historic range have declined considerably in condition and extent in the last 30 years (Lamb et al., 1993; Rennison, 2016; Rohweder, 2000; Stone et al., 2018). The aim of this study was to better understand the links between fire regimes and the condition and extent of the grassy sclerophyll forest habitat of the northern bristlebird. Using 30 years of fire history data, we model the degree to which fire histories correlate with northern bristlebird presence, as well as habitat extent, vegetation structure and invertebrate availability within their grassy sclerophyll forest habitat. Specifically, we ask: 1) is the persistence of northern bristlebirds related to the fire history of a site? 2) are grassy habitat extent (‘patch’ size) and understorey vegetation structure within patches associated with particular fire regimes? and 3) is prey availability associated with the fire history of a site?

5.3. Methods

5.3.1. Study area

Fieldwork was carried out at all 38 known currently and historically occupied northern bristlebird sites (a group of territories that occur close together, defined in Chapter 2 page 28 and Fig. 3.1.). These sites are dispersed across south-east Queensland and northern New South Wales, with regional clusters found in and around Conondale, Main Range, Lamington and Border Ranges national parks (Fig. 5.1). Sites are situated on the subtropical-temperate (Stern et al., 2000) climatic boundary and experience warm, wet summers and mild, dry winters. Rainfall averages 947 mm/year, most of which falls during the summer, and air temperatures average 32 °C in summer and 22-26 °C in winter (Bureau of Meteorology, 2014). Territories that are still occupied by bristlebirds are now

mostly found on private property adjacent to the parks, and are mainly on upper slopes or exposed ridgelines where rainforest gives way to open sclerophyll habitat. Much of the open sclerophyll woodland found on lower slopes in the study region has been cleared for agriculture. Rainforest is the dominant vegetation at higher elevations. Both cleared land and rainforest are uninhabitable by bristlebirds, resulting in limited connectivity between habitat patches.

a 1

1981 b

Brisbane

2009 c 2 3 4

5.3.2. Northern bristlebird presence/absence

Northern bristlebird presence and absence was determined using data from biennial monitoring of all known territories, as determined by Holmes (1989) (Sandpiper Ecological Surveys, 2012; Wildsearch Environmental Services, 2014, 2016b). Regular monitoring of these since 1999 using consistent methodology and personnel provided presence/absence data at each territory. Biennial monitoring consists of a five-minute broadcast of northern bristlebird recordings followed by a 20-25 minute listening period. A full description of the methods can be found in Stone et al. (2018). Since 2013, biennial monitoring has been supplemented by use of a specialised detector dog owned by the NSW Office of Environment and Heritage that is capable of detecting bristlebird material (fresh, aged or diluted feather and scat samples) to an accuracy of within 2 m of target (L. Baker pers. comm.). This has provided high detection accuracy for the usually cryptic northern bristlebird and has been used to confirm absences at historical sites. Bristlebirds were classified as absent when no sign of them had been recorded within the last five years. Of the 38 sites included in this study, 16 were occupied. The remaining 22 had been abandoned since the original survey of Holmes in 1982. Sites were delineated based on the territory mapping of Holmes (1989), and – for newer sites that had not been recorded by Holmes – on territory identified in biennial monitoring surveys (Sandpiper Ecological Surveys, 2012; Wildsearch Environmental Services, 2014, 2016b).

5.3.3. Fire history

Fire history data from a range of sources were collated for all 38 sites (Appendix 4) for the period 1980 to 2016. These included NPWS/OEH official fire extent mapping for national parks in NSW, reports by private monitoring contractors, historical site reports and property management strategies, supplemented by consultation with local landholders, consultants and land managers. We were able to find reliable dates for fire events for each site back until the early 1980s thanks to detailed fire records that were kept by stakeholders. Prior to the 1980s, fire records are less reliable, particularly on private land, therefore we set our timeframe for 1980 to 2016 based on the high confidence in records available during this period. From this dataset, we calculated: (i) time since most recent fire; (ii) mean fire interval, and (iii) longest recorded interval between any two fires. These variables were chosen because the literature suggested of the different components of fire

regimes, these would have the strongest effect on vegetation composition and structure in our sites (e.g. Cary & Morrison 1995; Morrison et al 1995; Watson & Wardell-Johnson 2004; Gosper et al 2012, Tasker & Watson 2017).

5.3.4. Vegetation characteristics

Four to six 20 x 20 m survey plots (depending on site size) were randomly placed within 200 m of known northern bristlebird locations at all 38 northern bristlebird sites while invertebrate samples could only be collected at 23 sites (occupied = 9, absent = 14) due to accessibility constraints (see Chapter 4).

At each site, understorey vegetation structure was assessed in May-July 2014 and 2015. Grass height was measured along the two central axes of a 20 x 20 m plot (two x 20 m transects) at 50 cm intervals, and averaged for the site (i.e. plots were lumped). Total grass and graminoid cover (%) was estimated within each 10 x 10 m subplot and averaged across all plots at the site to provide an assessment of ground cover. In addition, individual cover estimates for each grass and graminoid species were made within each subplot and then averaged for the entire site. Tussock cover was calculated as the total cover of clumping grass species (Poa, Sorghum and Themeda spp.). For more detail on methods used to measure the full suite of vegetation characteristics refer to Stone et al. (2018).

5.3.5. Invertebrate resources

Invertebrate samples were collected from each survey plot in 2015 and 2016 in May-June and September-November, which correspond respectively to non-breeding and peak breeding seasons for the northern bristlebird. Invertebrates were collected using pitfall traps and leaf litter samples. Each survey plot was divided into sixteen 5 x 5 m subplots. A single pitfall trap was placed centrally in one randomly selected subplot (a different one on each sampling occasion), while the single leaf litter sample was collected in a different randomly selected subplot to avoid interactions between the sampling methods (Ward et al., 2001a). Pitfall traps were left open for seven nights on each occasion. The traps were filled with approximately 3 cm of a diluted propylene glycol fluid and detergent. When pitfall samples were collected, the leaf litter samples were collected by gathering leaves, litter and loose humus from a 50 cm2 quadrat dropped in the centre of the pre-determined

subplot. Leaf litter samples were stored in fine-weave cotton bags until sorting. This was done using a Berlese extraction apparatus for one week. All invertebrate samples were then stored in 70% ethanol and sorted to order. Once sorted, the dry weight (in grams) of each sample was calculated by drying in a 60° C oven for 48 hrs. For a full description of invertebrate collection methods refer to Chapter 4.

5.3.6. Habitat patch extent and canopy cover

The current extent of habitat was mapped by visually delineating the complete contiguous patch of grassy forest that encompassed known (historic and current) northern bristlebird sites using ADS40 stereo-imagery, which has a resolution of 5 m, allowing fine scale mapping at a scale of less than 1:2,000. The most recent available imagery was from 2009. The historic extent of habitat was mapped by visually delineating patches of grassy forest at the same northern bristlebird sites as above, but using black and white aerial stereo-imagery taken in 1981 at a scale of 1:25,000. Mapping was undertaken at a constant scale of 1:5,000 for both image sets to minimise possible confounding from the different image resolution, as the fine scale detail discernible in ADS40 imagery was not always as clear in 1981 imagery if a higher resolution was used. The aim of the mapping was to delineate the complete boundaries of contiguous grassy forest encompassing and surrounding historic and currently occupied northern bristlebird sites. In practice, these areas were readily discernible on both the 2009 and 1981 imagery.

Northern bristlebird habitat was identified based on an amalgamation of fine scale habitat classes used by Rennison (2016) during a concurrent project that mapped changes in northern bristlebird habitat between 1966 and 2009 for New South Wales. For our study, which covered the entire range of the northern bristlebird, we needed imagery that covered both Queensland and New South Wales: comparable imagery was only available for 1981 and 2009. As 1981 and 2009 imagery differed in their quality and resolution, we classified bristlebird habitat as follows (Fig. 5.2):

1) Understorey: Dominance of tussock grasses in the understorey. Field visits and historic monitoring reports showed that tussocks mainly consisted of Sorghum leiocladum, Poa siberiana and Themeda australis. Tussock grasses could be

visually identified in imagery by browner colouration (a) and clumping (b), compared to the uniform texture of exotic or pasture grasses; 2) Canopy: An open canopy that allows substantial light penetration to the ground. Eucalyptus canopy can be distinguished from rainforest canopy by the difference in green-grey vs green colouration, and crown shape and texture. Open woodland/woodland versus sclerophyll forest could then be separated based on the spacing between individual tree canopies. Open woodland is defined as having a projected foliage cover of 5 - 10%, whilst woodland has a projected foliage cover of 10 – 30% (Specht, 1970); 3) Mid-stratum: A sparse mid-stratum with few shrubs and weeds; e.g., Lantana which is distinguishable as a thick, bright green layer (c) or shrub thickening which can be identified from loss of visible grass layer (d); and 4) Close proximity (less than 1 km) to rainforest.

a Woodland Sclerophyll forest Rainforest c

b d

Canopy gaps are an important attribute of northern bristlebird breeding locations, with all known nests built in open locations with no canopy cover directly overhead (D. Charley & A. Beutel pers. comm.). Once polygons were drawn around distinct grassy habitat patches, supervised classifications of canopy cover were conducted within each polygon

for both 1981 and 2009 aerial imagery in ArcMap based on methods detailed in Ibanez et al. (2013). Each individual pixel within a polygon was assigned as either open ground, tree canopy or shadow (which could not be clearly identified as either open ground or canopy cover) based on manually selected training areas. Training areas-small polygons used to guide classification-were drawn for each class allowing for a minimum of 500 pixels within each site polygon to be manually identified (larger patches required higher coverage). Once classifications were run across each polygon, their validity was then checked using 492 randomly sampled classification points (246 open ground, 246 canopy cover) and compared to manual classifications. We found over 85% of the classified points were correctly classified under the model.

5.3.7. Statistical analysis

Change in habitat/canopy extent

To examine how the extent of grassy habitat patches has changed between 1981 and 2009, we conducted paired t-tests on both the extent of grassy habitat and canopy cover across all northern bristlebird sites between 1981 and 2009. We then modelled how the amount of change (%) for habitat extent and canopy cover predicted northern bristlebird persistence.

Northern bristlebird presence

We used an information-theoretic approach to assess the relative support for two alternative models, fire history and change in habitat extent, in predicting northern bristlebird persistence using alternative generalized linear mixed-effects model (GLMMs). Both null and alternative models were fitted with a logit link using the lme4 R package (R Development Core Team, 2014). We controlled for spatial auto-correlation of northern bristlebird sites by incorporating an appropriate auto-covariate in both models (Crase et al., 2012) and standardised all variables. We compared the null model with a global fire history model to test for the effect of time since most recent fire (TSF), mean fire interval (MFI) and longest known inter-fire interval (LFI) in predicting northern bristlebird presence. Our habitat change model tested the effect of change in habitat extent (CHE) and canopy cover (CCC) on northern bristlebird persistence.

Habitat condition

Understorey vegetation structure, particularly grass structure, plays an important role in northern bristlebird habitat suitability (Rohweder, 2000, 2006; Stone et al., 2018). To determine how this resource is influenced by fire history of bristlebird territories, we used GLMMs to model grass structural variables against fire history. The grass variables used for this analysis were selected based on their importance for the northern bristlebird (Stone et al., 2018). We modelled mean grass volume (height x cover), mean tussock cover and percent of dominant grass species cover using a global fire history model that included time since fire, mean fire interval and longest known inter-fire interval as predictors. In addition to vegetation structure we wanted to know whether food availability may also be influenced by fire history in a site. Invertebrate biomass was modelled against time since fire, mean fire interval and longest known fire interval.

5.4. Results

5.4.1. Change in habitat extent between 1981 and 2009

The average size of grassy habitat patches reduced significantly from 1981 to 2009 (t = - 5.108, df = 43, p = <0.001). In 1981, the overall extent of available grassy habitat patches (that contained northern bristlebird sites) averaged 337.99 hectares (± 158.8), while by 2009 they had declined to an average of 152.07 hectares (± 87.9). By 2009 available grassy forest habitat for northern bristlebird had more than halved, declining to 44% of its 1981 extent. While habitat extent significantly differed between 1981 and 2009, no difference in canopy cover within sites was observed, indicating that the site boundaries were experiencing tree and shrub encroachment rather than the interior of the sites (Fig. 5.3).

5.4.2. Role of fire history and change in habitat extent on northern bristlebird presence

Fire history was the best performing model for predicting northern bristlebird presence, outperforming the null and habitat change models (Table 5.1). Northern bristlebird presence was more likely in sites that were more recently burnt, had shorter mean intervals between fires, and whose longest interval was relatively short. Sites with even a single long fire interval were less likely to have northern bristlebirds.

Table 5-1. Relative performance of alternative models for northern bristlebird persistence in response to fire history and change in habitat extent between 1981 and 2009.

Model coefficients Time since Mean fire Longest fire Change in Change in Model df logLik AICc delta fire interval interval habitat extent canopy cover -39.292 -0.101 -56.858 Fire history 6 -2.68 18.07 0 (± 0.90) (± 2.03) (± 3.50 ) Null 3 -7.02 20.94 2.87 -0.46 3.003 Habitat change 5 -4.47 20.82 2.94 (± 1.53) (± 1.86)

We found no difference in the amount of habitat lost between current and historically occupied bristlebird sites (Fig. 5.4), reflecting the fact that all sites had a similar level of habitat loss across the study region. Abandoned northern bristlebird sites had higher variance in the percent of change in canopy cover between 1981 and 2009, but this was not significant (f = 1.9483, df = 21, p = 0.1892). This may reflect that the canopies of some of these sites have been thinned or removed entirely, while other sites have changed from open forest to closed forest since 1981. Overall, we found no evidence that currently occupied sites have had less (or more) habitat loss or changes to canopy structure compared to historically occupied sites.

5.4.3. Influence of fire on habitat extent and condition

The substantial decline in extent of grassy habitat was strongly associated with fire history. We found that the longest known inter-fire interval had a strong influence on the percent change in the area of grassy forest in which sites were located between 1981 and 2009 (Table 5.2).

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Table 5-2. Relative performance of alternative models testing influence of fire history (1981- 2009) on grass volume, tussock cover, dominant grass cover, invertebrate biomass, and change in habitat extent within all bristlebird sites. Model coefficients are time since fire (TSF), mean fire interval (MFI) and longest inter-fire interval (LFI). Table compares AICc values for the null model with the AICc of the alternative fire history model. Model coefficients indicate the importance of each fire history variable within the fire history model.

Null model (df = 3) Fire history model (df = 6) Model coefficients logLik AICc delta logLik AICc delta TSF MFI LFI Grass volume -331.2 669.1 37.79 -308.3 631.32 0 -363.73 -353.46 -210.01 Tussock cover -121.2 248.2 3.13 -116.6 245.07 0 0.004 -0.083 -0.045 Dominant grass cover -148.64 303.99 5.76 -141.76 298.23 0 2.939 -4.427 2.217 Invertebrate biomass 5.146 -3.464 0 -0.95 17.13 20.6 -0.03 0.05 0.004 Habitat change -190.92 388.55 10.59 -181.63 377.96 0 2.45 -7.45 9.01

Of all three fire history variables, longest fire interval had the strongest effect on the amount of habitat change at northern bristlebird sites. Decline in habitat extent was more severe when extended fire-free periods occurred. For example, habitat patches that had no significant change in habitat extent (n = 4) all had a longest known fire interval < 10 years. Mean fire interval also had a strong effect: the few patches that did increase in grassy habitat extent (n = 3) all had mean fire intervals of < 8 years. The mean fire interval for all sites between 1981 and 2016 is 8.54 years (8.4 for currently occupied sites), with an average longest fire interval of 15.75.

Habitat condition was also strongly influenced by fire history at northern bristlebird sites (Table 5.2). Grass volume, tussock cover and dominant grass cover were all negatively related to mean fire interval. Mean fire interval had the strongest effect on tussock cover and dominant grass cover, while grass volume was most affected by time since fire and mean fire interval (Fig. 5.5). Sites with longer mean fire intervals and a long longest inter- fire interval had a lower volume of grass. While grass volume declined with increasing time since fire, the percentage cover of the dominant grass species at a site increased with increasing time since fire, suggesting a decrease in the diversity of grasses over time (Fig. 5.5). Sites with a mean fire interval of < 3 years (n = 4) had reduced grass height and grass diversity, 10% less tussock cover (particular of Sorghum), and a 20% increase in dominant grass (T. triandra) cover.

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a b

Figure 5.5. The relationship between time since fire on (a) grass volume (height x cover), and (b) cover of the dominant grass species (%) in all bristlebird sites.

In contrast to both bristlebird presence and vegetation structure, we detected no effect of fire history on invertebrate biomass (Table 5.2).

5.5. Discussion

The extent of grassy habitat known to have been used by northern bristlebirds had declined considerably between 1981 and 2009. This decline related to changes in fire regimes, with longer periods of fire absence associated with decreasing availability and condition of grassy habitat. Longer mean fire intervals also were associated with changed understorey structure and condition of remaining grassy areas, habitat attributes that have been shown to be vital to northern bristlebird persistence (Stone et al., 2018). Grass volume and tussock cover were lower at sites with longer fire intervals and where a prolonged period of fire absence has occurred. Northern bristlebirds themselves were more likely to occur at sites that had more frequent fire with shorter fire intervals and short periods of fire absence.

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5.5.1. Contraction of grassy patches

Mapping of open grassy sclerophyll forest across all current and historically occupied northern bristlebird sites showed a drastic decline in available grassy habitat between 1981 and 2009. During this 30-year period, extent declined by roughly half. Open forest or savannahs located adjacent to rainforest appear to be highly susceptible to rainforest expansion, not only in Australia (e.g. Bowman 2000, Stanton et al 2014) but all over the world (Murphy & Bowman, 2012; Tng et al., 2012). Our results demonstrate this is also the case in northern bristlebird habitat in areas adjacent to eastern Australia’s Gondwanan rainforests. As Stone et al. (2018) demonstrated that northern bristlebird presence is strongly associated with large patches of grassy eucalypt forest with tall thick grass structure, the significant reduction in extent of such habitat is of concern.

Northern bristlebirds are territorial, and individual home ranges vary between 2-5 hectares depending on bird density (Holmes, 1989; D. Charley pers. comm.). In a review of the minimum viable population size literature since 1980, Frankham et al. (2014) suggested that as a general rule across taxa a minimum population size of 100 is necessary to avoid inbreeding depression, and 1000 individuals are needed to retain the long-term evolutionary potential of a species. Given that only three occupied northern bristlebird habitat patches are larger than 100 hectares (which hold 8 sites), and the rest average < 60 ha in size, the carrying capacity of remaining patches might be below that required for a viable population. Therefore, reclaiming now-marginal ecotone habitat to increase the core area of grassy habitat extent is critical for the future persistence and survival of the northern bristlebird.

In addition to genetic pressures, populations in small, fragmented habitat patches often have increased risk of local extinction from stochastic events (Clark et al., 1989; Shaffer, 1981; Wilcox, 1986). While we show that frequent fire is necessary for maintaining suitable habitat, fire events themselves can directly cause the loss of small and isolated subpopulations (Hanski et al., 1995), particularly if they affect an entire patch and affected animals are unable to relocate to suitable habitat. Small, degraded habitat patches may be less likely to burn as often because they have less continuous ground fuel and moister leaf litter due to shading , but in extreme weather when conditions enable flames to connect more patchy fuels (Bradstock, 2010; Zylstra et al., 2016) the result can be catastrophically

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severe fires. Frequent, low intensity fires in grassy patches are likely to increase habitat heterogeneity by maintaining their difference in structure from adjacent rainforests.

Change in habitat extent between 1981 and 2009 was strongly influenced by fire history, and particularly by the longest known fire interval in that period. Northern bristlebirds live in a highly productive landscape: the high rainfall and basalt soils result in high plant diversity in a competitive environment (Low, 2011). Several rainforest plant species can replace grassy habitats in the absence of fire (Butler et al., 2014; Chapman & Harrington, 1997; Unwin, 1989). Shrub encroachment following reduced fire occurrence has been well documented in grassland and savanna (Blanco et al., 2014; Bond, 2008; Bowman & Fensham, 1991; Fensham & Fairfax, 1996; Hoffmann et al., 2012b; Williams et al., 1999) but less attention has been given to encroachment into natural grassy forested habitats which are often viewed as degraded forest systems (Parr et al., 2014). Our findings suggest that even a single period of extended fire absence may have provided the opportunity for shrub establishment and have accelerated rainforest encroachment into grassy habitats (Burgess et al., 2015). Common mid-storey plant species such as many Acacia produce large quantities of seed once mature and can germinate en masse following hot fires (Gordon et al., 2017). In and adjacent to northern bristlebird sites, Acacia melanoxylon and A. implexa are common mid-stratum species, but frequent low intensity fires may limit seed germination and seedling establishment by not being hot enough to break seed dormancy, but able to kill small seedlings, thus maintaining a ‘fire trap’ (Grady & Hoffmann, 2012; Hoffmann, 1999; Werner, 2012).

5.5.2. Altered habitat condition

Grass volume and tussock cover were strongly influenced by fire history, with reduced grass volume and tussock cover at sites that had longer mean fire intervals and longer fire- free (i.e. longest inter-fire) periods. Following fire, grass cover can recover quickly and observations at northern bristlebird sites suggest grass volume is greatest one to two years post-fire (P. Watson, D. Charley pers. comm.). Tall, thick grasses are an important habitat component for the northern bristlebird (Holmes, 1989; Lamb et al., 1993; Rohweder, 2000; Stone et al., 2018). Height and sward structure of tussock grasses is known to decline with increasing time since fire (Morgan & Lunt, 1999), which in turn can lead to tussock collapse and consolidation, and thus a decline in refuge and nesting

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resources. The northern bristlebird is highly terrestrial, and spends most of its time on foot on the ground in a manner that has been likened to a small terrestrial mammal, (‘a rat in a feathered suit’, Ford Kristo 1997 in Baker (1998a)), turning over leaf litter with its beak. Short fire intervals which reduce compaction of the grass layer and promote healthy grass growth are apparently essential for northern bristlebird persistence.

The cover of grass tussocks showed a weaker response to time since fire than other grass variables, which may be related to both their quick growth rate and our coarse classification of time since fire into years. Tussock species often do not completely burn in a low intensity fire, and are capable of re-sprouting extremely quickly following fire (P. Watson pers. comm.). Mean fire interval and longest fire interval had a stronger influence on tussock cover. Increases in inter-fire intervals may contribute to loss of tussock through causing reduction in sward health and seedling recruitment at sites. McDougall (1989) showed that older tussocks at infrequently burnt sites had lower seedling recruitment due to a build-up of dense compressed material at their base. Extended fire absence may also reduce the ability of large tussocks to regenerate following fire. Compression of grasses reduces light available for tussock growth, or seed germination, and may lead to eventual replacement of tussock species (Morgan & Lunt, 1999). While non-tussock grasses may provide acceptable conditions for individual survival, ongoing loss of tussock species from grassy sclerophyll forest patches may be affecting the breeding success and long-term persistence of northern bristlebirds because of the dependence on tussocks for nesting. In addition, taller tussock grasses on the rainforest margin may provide increased invertebrate resources (Stone et al. (2018), thus a loss of tussocks may affect food resources for the northern bristlebird as well as safe foraging and nesting micro-habitats.

5.5.3. Conservation implications

Our research shows that reduced fire occurrence in this landscape has contributed substantially to the decline of grassy sclerophyll forest habitat for the northern bristlebird. Northern bristlebird habitat is adjacent to rainforest, and I show it needs relatively frequent fire to maintain grass and tussock cover in a competitive rainforest landscape. A bias towards rainforest protection in our study region (Shoo et al., 2014) may be contributing to the decline of grassy forest patches that occur within the rainforest matrix. Today, currently

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occupied northern bristlebird sites are located predominantly on private property (10 out of the 11 currently occupied sites) where landowners have continued burning largely for grazing management purposes. Regular ecological burns will be important for maintaining this grassy forest habitat. Patches should not remain unburnt for more than eight years if we are to avoid shrub encroachment and habitat loss. As such, land managers should develop appropriate fire regimes through tracking fire frequency and planning burns based on this information. Increased resourcing and facilitation for ecological burns and awareness of their value to grassy forest habitat will be key, particularly for land managers that are limited in skills or resources for carrying out ecological burns.

Habitat loss appears to be a significant, if not key, contributor to the decline of the northern bristlebird. Our findings suggest that fire intervals of greater than 8-10 years have contributed to the vegetation changing to a point where it becomes resistant to fire. Considering the average fire interval 8.4 at occupied sites, active habitat management and prescribed burning is needed to reduce the fire interval and avoid sites reaching a fire resistant threshold. Without frequent fire, rapid shrub encroachment can occur and grassy habitat declines. Northern bristlebirds, and several other threatened species, are likely to become locally extinct without active management of fire to maintain these open grassy forests.

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Chapter 6

Thesis conclusion & recommendations

Plate 5. Captive northern bristlebird singing at Currumbin Wildlife Sanctuary.

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6.1. Thesis overview

In the research undertaken for this thesis I have investigated the influence of habitat structure, resource availability and disturbance regimes on the persistence of the critically endangered northern subpopulation of the eastern bristlebird. I have also extended knowledge of how fire regimes regulate the extent and understorey composition of grassy sclerophyll forests in a highly productive landscape dominated by rainforest within the south-east Queensland bioregion of Australia. The conservation importance of these forests has been largely overlooked. The main objectives of the research conducted for this thesis were to:

 build understanding of the relative roles of vegetation structure, invertebrate prey and the landscape context of grassy sclerophyll forest habitat patches in determining habitat suitability for the northern bristlebird, and  determine the role of fire in maintaining extent and condition of grassy sclerophyll forest habitat and its resources for northern bristlebirds.

These objectives were addressed throughout Chapters 2 to 5 by answering the following specific research questions:

What is the current state of knowledge of northern bristlebird ecology and habitat dynamics? Which structural characteristics of grassy eucalypt forest habitat are associated with northern bristlebird persistence? Does food resource availability differ between habitat patches in which northern bristlebirds have persisted and those in which they have not? How does fire history influence the suitability of habitat patches for northern bristlebirds, and what fire regimes are most appropriate to maintain their habitat?

Question 1 was addressed in Chapter 2, where I reviewed the existing state of knowledge of northern bristlebird ecology and habitat requirements. While no published information was hitherto available on the northern bristlebird, there is a wealth of knowledge held by the Northern Working Group and individual experts that have been involved in the conservation of the bird for over 20 years. I collated this information and used it to

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summarise the characteristics of northern bristlebird habitat and their ecology. I then developed a conceptual model that represented a testable hypothesis of the mechanisms that control habitat suitability for long-term northern bristlebird persistence.

Previous monitoring of northern bristlebird and their habitat by Holmes (1989), Lamb et al. (1993), and Hartley and Kikkawa (1994) all demonstrated that birds favoured areas of tall, thick grassy eucalypt forest along the rainforest margin. Since then, northern bristlebirds have been lost from many of their original territories, and their current pattern of occupation may be limited by the specific habitat characteristics present within sites. For Chapter 3, I was able to develop this hypothesis in Question 2: testing whether northern bristlebird persistence in known sites related to specific habitat characteristics. To answer this, I surveyed all known historic and currently occupied northern bristlebird sites (excluding the few historical sites that were in heath, and sites that no longer contained any grassy habitat) to see whether fine scale differences in vegetation structure or habitat patch context correlated with where northern bristlebirds remain. I compared the explanatory power of alternative models of vegetation and patch context in predicting current northern bristlebird presence. I found their continued presence was strongly associated with the extent of grassy habitat patches. Large habitat patches were more likely to have northern bristlebirds than smaller patches. This relationship was, however, contingent on the condition of the grassy understorey, particularly grass height and grass cover. Habitat patches with a taller, thicker grass layer had a higher probability of northern bristlebirds occurring. Large habitat patches without this understorey structure had a lower probability of containing northern bristlebirds. I found little evidence that differences in species composition of the vegetation played a role in northern bristlebird persistence.

Before this study, no research had been conducted on northern bristlebird food availability within their grassy habitats to see whether this might be a contributor to their decline. Gibson and Baker (2004) observed foraging and feeding behaviour of southern bristlebirds in their heath habitat and found invertebrates were the dominant component of their diet. Because of this, I examined whether invertebrate resources may be limiting the persistence of northern bristlebirds. To do this, I collected invertebrate samples using pitfall and leaf-litter trapping at all historic and currently occupied sites to detect if there

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were differences in food availability. As presented in Chapter 4, I found no difference in the composition or quality (abundance, nutritional value or biomass) of invertebrates. I concluded that northern bristlebird presence was not associated with invertebrate density, nutritional value or composition. However, there were a number of limitations to this study. Variation in invertebrate availability may not have been identified due to only autumn and spring seasons being sampled, natural variability of invertebrates across years and collection biases towards small, ground dwelling invertebrates. Food limitations may be more important during winter or summer when conditions are more extreme, compared to breeding periods. In addition, seed resources may be a stronger limiting factor then invertebrates, providing important nutrients that are obtained through invertebrate diets. Despite these limitations, it seems likely that variation in prey availability is not driving northern bristlebird persistence. This may be because of the generalist nature of northern bristlebirds and the high abundance of invertebrates I collected at both current and abandoned sites.

Finally, in Chapter 5, I presented the results of my fourth research question: to determine how historical fire regimes have influenced the habitat and persistence of northern bristlebird. Previous research on fire and the northern bristlebird has been fairly limited, and has focussed on individual fire events, and investigated vegetation recovery following wildfires at a few locations (Lamb et al., 1993). By comparison, the response of the southern bristlebird population to fire, particularly around Jervis Bay and Booderee National Park, has received relatively in-depth attention (Bain et al., 2008; Bain & McPhee, 2005; Baker, 1997, 2000; Lindenmayer et al., 2009; Pyke et al., 1995). The research on the southern population and its heath habitat has largely dominated the thinking about the fire ecology of the species, and as I show may be resulting in detrimental outcomes for the northern subspecies and its very different habitat.

I collated 30 years of fire history information for the 38 historical and currently occupied northern bristlebird sites where grassy habitat occurs. Since 1989, the northern bristlebird population at these locations has declined by over 80% (see Chapter 2). I found that over the same period, the size of remaining grassy habitat patches has decreased by more than 50%. I tested whether this decrease was related to site specific fire histories of the habitat patches. Fire history had a strong influence on habitat loss, with longest fire interval

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and mean fire interval having the greatest impact on amount of habitat lost and current habitat condition. Sites with longer periods of no fire (longest known fire interval) and longer mean fire intervals had greater rates of habitat decline. Given the findings in Chapter 3 that northern bristlebird persistence was strongly associated with habitat extent, the substantial decline in habitat across the landscape is concerning. To test whether current northern bristlebird presence itself was associated with habitat decline and the fire history of a site, I carried out statistical analyses and found that this was the case. Longest fire interval was the strongest predictor of northern bristlebird absence. This result is consistent with the hypothesis that long periods of fire absence, and the associated decline in habitat extent, are substantial contributors to the severe decline of the population over the past three decades. Based on these findings, as presented in Chapter 5, I concluded that regular fire is a vital part of maintaining habitat for northern bristlebirds. This is in contrast to the situation for southern subpopulation, and thus the published literature on the species to date has not emphasized the impact that disappearing fire may be having within northern habitat.

By addressing these four main research questions in this thesis, I have discovered valuable information for the management and recovery of the critically endangered northern bristlebird. At the end of this synthesis, following a series of sections in which I outline the contribution my study has made to a range of disciplines, I draw all of these strands together to provide management recommendations for the northern bristlebird, and address my final research question:

Based on evidence procured throughout this thesis, what is the best management strategy for conserving grassy forest habitat?

6.1. Contributions of this thesis to scientific knowledge

This thesis makes a number of valuable contributions to understanding the ecology and habitat requirements of the previously little-studied, but highly threatened, northern bristlebird. It also advances our understanding of grassy forests, fire ecology and the understanding of critical habitat elements and their dynamics that can help inform management actions for threatened species. Although my thesis focussed on a single

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species, the findings have broader implications for research on drivers of species persistence, and understanding the causes of decline and persistence in modified habitats and disturbance regimes. My research has highlighted the importance of habitat assessment prior to release for reintroduction programs that aims to understand both the causes of decline in the wild as well as what is needed for the species to be able to persist. Without such assessment, reintroduction is unlikely to adequately predict what factors are critical for the survival of released birds, and thereby maximise population recovery and persistence.

6.1.1. Fire ecology and vegetation dynamics

Loss of forests and woodlands from land clearing has dominated much of the research on conservation outcomes in forested landscapes (Reside et al., 2017). However habitat modification from altered disturbance regimes can also have significant impacts on biodiversity. Parr et al. (2014) demonstrated that modification of tropical grassy biomes is an underestimated challenge for biodiversity. Fire is a crucial element in determining the location of grassland-forest boundaries, and the alterations of the natural, or pre-European fire regime can have serious implications for forest and shrub encroachment, and alter grassland characteristics (Geiger et al., 2011; Ibanez et al., 2013; Sharma et al., 2014). Recent research has shown how mobile the grassland-forest boundary is, and that changes to fire regimes can lead to long-term consequences for vegetation communities (Blanco et al., 2014; Breman et al., 2012; Butler et al., 2014; Eldridge et al., 2011; Fairfax et al., 2009a; Geiger et al., 2011; Hoffmann et al., 2012a; Hoffmann et al., 2012b; Just et al., 2016; Kirkman et al., 2014; Moxham et al., 2016). These studies, however, have mostly focussed on grassland or savanna, and the importance of fire for maintaining forested grassy habitats has received less attention (Ratnam et al., 2011). My research has contributed towards filling this gap, providing an example of how grassy forests can provide important resources for a threatened species, and how critical frequent fire is for maintaining them in areas of high productivity.

Grassy forests are an ecotonal habitat often found between intact closed forests and cleared agriculture land, leading them to be often regarded as degraded or human modified habitat (Bond & Parr, 2010; Ratnam et al., 2011). I have shown that grassy

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forests are an important habitat in their own right, and because of their ecotonal nature are possibly more sensitive to modified fire regimes than more open grasslands and savannas (Luza et al., 2014). Hoffmann et al. (2012a) characterised grassy systems that are left unburnt as able to reach a ‘fire suppression threshold’, that is a state in which they are less likely to burn when canopy cover increases to a point where it suppresses grass growth, and thus the continuous near-surface fuel that readily carries frequent fire. I showed that this potential fire suppression threshold (i.e. loss of grass layer) may be achieved rapidly during a single extended period of fire absence, supporting the importance of maintaining frequent fire in grassy forest habitats. I suggest that limited knowledge of the importance and habitat dynamics of grassy forests within rainforest orientated thinking by land managers in the past may have contributed to their decline through encroachment from adjacent rainforest, and the concurrent decline in the threatened species that rely on them.

Understanding the feedbacks between vegetation and fire is important for the management of grassy habitats, particularly for those that exist alongside closed forests (Bond & Parr, 2010; Hoffmann et al., 2012b; Ratnam et al., 2011). I provide evidence that the period in which fire suppression (see Fig 6.1) is reached in high elevation grassy forests may be quick, and transitions to rainforest can occur within a decade if regular fires are not maintained. This has implications for the management of grassy sclerophyll forests and maintenance of a healthy rainforest-grassy forest mosaic within the globally recognised Gondwanan Rainforest World Heritage Area (UNESCO, 2017).

6.1.2. Adaptive management of grassy forests

Requirements for grassy forest management are likely to vary greatly across sites depending on environmental conditions and topography. Management of this habitat will therefore rely on an adaptive approach for maintenance and restoration of it across the landscape. Variability in topography and environmental conditions means that the timing, extent and intensity of fires needed to maintain or improve habitat condition will vary, therefore land managers will need burn habitat through an adaptive fire management approach. Adaptive management has become a central idea to biodiversity conservation, supporting re-evaluation of management strategies based on monitoring and on ground responses (Lindenmayer & Burman, 2005). For grassy forests, active adaptive

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management based on on-ground decisions will be essential to ensuring grassy habitat does not transition to rainforest. This form of on-ground, real-time decision making for grassy forest management was a strong part of aboriginal burning practises throughout Australia. In Australia, indigenous burning practises were highly influenced by local conditions and a deep understanding of their ecosystems. Indigenous knowledge of ecosystems, such as aboriginal burning practices, has been described as a form of adaptive management (Berkes et al., 2000). In this thesis I have demonstrated the importance of frequent fires for promoting grassy forests and how quickly vegetation structure can change when fire is absent from the landscape. I have also shown that this ecotonal habitat provides important resources for the northern bristlebird. The persistence of this species (and others) will depend on an adaptive approach to vegetation and fire management. The habitat model provided in this chapter (page 121) is a useful tool for land managers to identify management needs within grassy forest patches. While site variability may mean time frames, and specific vegetation responses will vary, this model can help land managers identify which stage their sites are at, and adjust management accordingly. Not only restoration of fire, but restoration of the adaptive aboriginal fire practises needs to be encouraged to help manage grassy forests appropriately.

6.1.3. Invertebrate ecology

My research added to our understanding of invertebrate ecology by investigating whether invertebrate communities in grassy sclerophyll forest varied in response to fire regimes and vegetation structure. While I found little evidence of a relationship between northern bristlebird persistence and invertebrate availability, there were several findings that suggested that invertebrate assemblages varied with habitat structure and context.

While I did not specifically test how invertebrates varied across the rainforest-grassy forest ecotone, I found higher nutritional value of large invertebrates closer to the rainforest margin. Most studies that have looked at invertebrate communities along a forest edge have focussed on boundaries with highly disturbed agricultural land (Kotze & Samways, 1999, 2001; Lacasella et al., 2015; Yu et al., 2007) or those that occur between forest and open grassland/savanna ecosystems (Pinheiro et al., 2010; Sharma et al., 2014). My findings of increased invertebrate quality closer to the rainforest margin suggest that the

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rainforest-grassy forest ecotone of northern bristlebird habitat may provide better conditions for nutritious invertebrates. The lack of variation in invertebrates between seasons also suggests the ecotone may provide more stable resources for insectivorous organisms through periods of environmental stress or resource bottlenecks and provide a buffer to environmental extremes which allow a more valuable invertebrate community to persist.

Food availability is a key determinant of species distributions; however, food availability for insectivores has received much less attention than other groups, particularly in regards to the nutritional value provided (Razeng & Watson, 2015). Common measures for assessing food availability include abundance and biomass, but for insectivores the nutritional quality of prey may be more relevant (Cosgrove, 2017; Razeng & Watson, 2015). I therefore tested the nutritional quality of invertebrates as a predictor of northern bristlebird presence in conjunction with the more common abundance and biomass measures. I found no evidence that either of these measures were predictors of northern bristlebird persistence. The lack of a significant relationship between invertebrate resources and northern bristlebird persistence suggests that structural changes to habitat at the landscape level (found in Chapters 3 and 5) may be more detrimental for some insectivorous birds, especially a generalist species such as the northern bristlebird, than sole measures of food availability (Cosgrove, 2017). It also suggests that less common dietary items such as seeds may play a more important role in northern bristlebirds than previously thought. The loss of tussocks from northern bristlebird habitat (as shown in Chapter 5) likely impacts the availability of this food source, which provides a different nutritional value to invertebrates.

Invertebrates receive far less conservation attention than other taxa. However, recent research has shown that globally, invertebrates are suffering high rates of decline (Hallmann et al., 2017). While I did not focus on invertebrate conservation in this thesis, I have collected a valuable invertebrate dataset that, through future comparative analysis, will contribute information on variation in invertebrate communities within this important grassy sclerophyll forest habitat. This dataset can also provide baseline information against which to compare future changes in prey availability, which may result from responses to habitat management actions or future environmental change.

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6.1.4. Reintroduction biology

Translocations and reintroductions are a common conservation action, and while rates of reintroductions have increased over the years, success rates have not (Fischer & Lindenmayer, 2000; Griffith et al., 1989). A commonly cited reason for reintroduction or translocation failures is inadequate information on habitat quality and threats at release sites (Wolf et al., 1998). The overall goal of this thesis was to contribute to assessment of habitat suitability for northern bristlebird reintroductions. In testing existing assumptions regarding habitat quality, I have identified the factors of greatest importance to northern bristlebird persistence, specifically habitat extent. I have also identified that reduced fire frequency in habitat patches is a key limiting process. This has important implications for bristlebird management, and it is hoped that this understanding will contribute to the successful reintroduction of northern bristlebirds and management of their habitat to ensure their long-term survival, a key element of reintroductions that is often overlooked (Taylor et al., 2017).

This research has provided an example of why it is critical to invest in quality pre-release research on habitat and thereby increase awareness of habitat suitability, species persistence patterns and threats prior to reintroduction attempts. While expert knowledge can act as a useful surrogate in the absence of quantitative data on habitat suitability prior to reintroductions, it is rarely validated (Cook et al., 2010). This can be a common feature within highly threatened species programmes, where the pressing need for on the ground management actions often trumps scientific research and publication of collected information (Linklater, 2003). I was able to formally test assumptions and hypotheses that have been developed through long-term monitoring of the northern bristlebird, but have not been verified by scientific analysis. I demonstrate how valuable this process can be in assessing habitat suitability for species persistence. There is often a wealth of information held within threatened species recovery teams that can be lost, or remain unknown to external researchers and managers, that may benefit from the information if readily available. Collating long-term monitoring data and expert knowledge, and formally testing hypotheses can be hugely beneficial not only to the individual species recovery, but to external groups as well (Burbidge et al., 2011).

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This thesis reinforces how different habitat requirements can be between populations of a given species, and how detrimental overlooking these differences can be to the long-term persistence of a population. I highlight how complex conservation within disturbance-prone habitats can be, and that understanding populations, particularly their site-specific habitat dynamics, is vital for determining appropriate management strategies (Landres et al., 1999).

6.1.5. Northern bristlebird ecology

To understand the conditions that will promote long-term persistence of northern bristlebird, we need to understand the causes of decline. During this thesis I have gathered information to assist with habitat management and future reintroduction of the critically endangered northern bristlebird. I helped test hypotheses from existing information and identify the most important factors associated with northern bristlebird persistence. This synthesis will help underpin a population-specific management and reintroduction strategy aimed at saving the northern bristlebird. The main findings presented in Chapters 2-4 relevant to northern bristlebird management are, in brief:

1. Northern bristlebirds occupy very different habitat to their southern counterparts and they therefore require population-specific management actions. 2. Northern bristlebird persistence is strongly associated with the extent and condition of grassy sclerophyll forest patches within the landscape. In particular, northern bristlebirds need large areas of grassy habitat, and appear more likely to persist if these grassy patches have tall, thick grass. 3. There was no discernible difference in invertebrate resources across northern bristlebird sites, leading to the conclusion that invertebrate composition, abundance or quality do not reflect current patterns of northern bristlebird presence. 4. While invertebrate resource availability appears to be relatively constant within the ecotonal grassy forest-rainforest margin that northern bristlebird occupy, fine-scale understorey structure does appear to influence the availability and distribution of invertebrates. 5. Grassy forest habitat patches around the QLD-NSW border that have recently supported bristlebirds have declined by > 50% in the last 30 years and this decline

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is associated with a reduced fire frequency and the associated rainforest/weed encroachment. 6. Northern bristlebird presence is strongly associated with frequent fire.

Based on these main findings, I provide below context-specific implications of this thesis for habitat management, conservation, and future reintroduction of the northern bristlebird.

6.2. Northern bristlebird management recommendations

6.2.1. Habitat management

Habitat management is already an important part of conservation efforts for the northern bristlebirds. Below I present specific habitat management recommendations drawn from the findings in each of the chapters in this thesis to help guide management decisions for maintaining suitable grassy habitat.

Chapter 3- Vegetation structure and habitat patch context

Based on my findings in Chapters 3, I recommend:

1. Large habitat patches are critical for long term survival of northern bristlebird. This is likely because of the territorial nature of the birds and their dependence on a highly dynamic habitat in which resources can be unevenly distributed. My research demonstrates that maintenance of patches of high quality, tussock dominated habitat at least 40 ha in size is vital for northern bristlebird persistence. Few large habitat patches remain (see Chapter 5), so the maintenance of these sites needs to be prioritised, especially where northern bristlebirds are still present in them. In addition, the size of remnant habitat patches should be increased (to 40-100 ha of tussock dominated habitat), tussock abundance improved within larger grassy habitat patches, and the connectivity between patches increased through appropriate fire and weed management. High quality small sites (of around 40 ha) will be valuable assets to northern bristlebird conservation only if connectivity is restored that allows birds to move between suitable patches as needed.

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2. Where possible, attempts should be made to restore patches of grassy habitat, particularly where rainforest encroachment has not completely converted habitat. Some landowners have been involved in manual manipulation of habitat trough shrub removal to help improve grass structure at their sites. Such experimental trials show that with intensive weed effort, grassy forest patches can be restored, however this is highly dependent on some grass structure still being present to provide a seed bank. Future research into how to restore tussocks to marginal habitat may be needed. 3. Northern bristlebirds prefer areas of grassy habitat in which the grass structure is tall (average height >40 cm), and thick and complex (cover of >100 % in the 0-1m vegetation stratum) (Chapter 3). When assessing sites for future releases, northern bristlebird land managers need to ensure that this thick, tall grassy understorey is present. I suggest that grass height and cover should be used as criteria for assessing site suitability. These criteria, in conjunction with patch size and connectivity to other habitat patches, are key for site suitability for long term northern bristlebird persistence. 4. A benefit of the 'classification tree' analysis in Chapter 3 is that it provides a small set of variables that are good indicators of likely northern bristlebird persistence. As part of the recovery efforts, the Northern Working Group conducts searches of historic habitat and new habitat to find new or relocated territories. Discovery of new territories would be of significant benefit to northern bristlebird conservation considering the small effective population size (5 breeding pairs) currently known. During field searches, rapid assessments of sites using my classification tree, in association with expert knowledge, may help direct subsequent searching for birds to areas that have the highest chance of having northern bristlebirds present. 5. While my research found no evidence that vegetation composition was important, a small number of tussock species do contribute much of the important structure required for northern bristlebirds. For instance, healthy S. leiocladum tussocks are tall in nature, and are one of the main species used for nesting by northern bristlebird. I suggest restoration and rehabilitation of tussock regrowth within existing habitat patches is an important management action.

Chapter 4: Invertebrate resources 119

Based on my findings in Chapter 4, I recommend:

6. Invertebrates are unlikely to be limiting the presence of northern bristlebirds. Because of this, management should focus on rehabilitating vegetation structure and habitat patch size. 7. Invertebrate availability was found to correspond to specific vegetation characteristics such as grass height and distance to rainforest margin. Taller grasses provide increased abundances of invertebrates, therefore the maintenance of these grassy conditions (also supported in Chapter 3) is recommended to promote food resources. 8. Experts within the recovery team have noted the strong relationship between northern bristlebird territories and the rainforest margin. Rainforest margins have previously been thought to provide refuge during fire events. In Chapter 4 I show that this grassy forest-rainforest ecotone also has more nutritious invertebrate resources. To maintain habitat that provides adequate resources for northern bristlebirds through periods of environmental stress such as drought, ensuring good quality grassy habitat on this ecotone remains intact may assist in the resilience of northern bristlebird to changing climatic conditions.

Chapter 5. Fire management

Based on my findings from Chapter 5, I have produced an updated habitat model of northern bristlebird habitat that will be useful for future management actions (Fig. 6.1). To accompany this updated habitat model, I recommend:

9. Fire intervals not exceeding 8-10 years are needed to conserve bristlebirds at sites where they exist. Sites that have had long periods of no fire (≥10 years) are likely to be difficult to return to optimal grassy habitat. My results showed that the period of longest fire absences was a stronger predictor for habitat loss than the mean fire interval, supporting the idea that such periods may accelerate rainforest encroachment into these grassy forest patches. This has been observed at some sites (e.g. Cougal Park) where recent prescribed burns have resulted in a higher abundance of shrubs germinating following the burn (Watson & Tasker unpublished data).

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10. Fire at intervals of less than three years may be detrimental to maintaining grassy structure that benefits northern bristlebirds. For the few sites that have a mean fire interval of < 3 years, reduced complexity of the grass community was observed, which may have negative impacts on invertebrate community as well (based on invertebrate relationships with grass structure found in Chapter 4). For example, northern bristlebird sites at Stretchers Track and Conondale Ranges have been subject to more frequent fire intervals (mean fire interval of 1.93 and 2.3 years respectively) which was correlated with reduced tussock cover and grass diversity, and an increase in dominance by T. triandra. 11. Little significant change in grass volume was observed in the first 5 years following fire, but began showing declines between 5-10 years following fire. This suggests that this may be the ideal time for prescribed burns to be implemented to maintain grassy habitat. Following 8-10 years without fire, there appears to be a stronger decline in grass height and cover and invertebrate abundance at sites, signifying a transitional period in which grassy forest habitat quality declines. As such, I recommend that fire intervals of 4-7 years should be implemented to maintain grass conditions suitable for northern bristlebirds. 12. Given the importance of large habitat patches for northern bristlebird, and the small overall size of sites currently remaining, prescribed burns at sites that are beginning to transition to a shrub dominated understorey should be used to increase extent of suitable grassy habitat. However, transitional sites will likely need a shorter fire interval for the follow-up fires to reduce shrub dominance. In some cases where shrubs have joined the canopy, thinning of mature shrubs may be needed to return canopy gaps which are necessary for healthy tussock growth. Even then, my findings suggest that tussock cover will be significantly reduced, and follow up remediation of sites may be required to re-establish critical tussock grasses for northern bristlebird breeding.

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1 1 • Large grassy habitat patches * Reseeding? • Grass height > 40 cm 3 • Ground cover 100 % or more (accumulative Fire interval total for 0-1 m strata) * <3 years • Higher abundance and nutritional value of invertebrates Homogenous, short Eucalyptus forest • Transition from stage 1 → 2 occurs within grass understory Tall grassy understory 2 approximately 8 years UNSUITABLE OPTIMAL • Grasses < 40 cm in height • Sparse homogenous grass layer (Blady grass) Long fire 2 *interval (>10) 3 • Biennial fires reduce tussocks and increase dominance of Themada triandra Fire interval <8 * years to avoid • Reduced grass diversity and potentially

FIRERESILIENCE transition to 4 invertebrate community Transitional threshold = 8-10 years no fire Eucalyptus forest * • Sparse grassy understory 4 Transition between 2 → 4 occurs between 8-10 years without fire MARGINAL 5 4 • Evident changes in vegetation structure towards closed forest • Fewer, larger invertebrates 10-20 years • Loss of tussocks if fire reintroduced (4 → 3) Rainforest Shrubby understorey • Irreversible transition to closed forest 10-20 5 UNSUITABLE TRANSITIONAL years without fire FIRESUPPRESSION Figure 6.1. Updated habitat dynamics model (from Chapter 2) for the northern bristlebird based on findings from Chapters 3-5 of this thesis. Stars denote information that has been added based on chapter results.

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6.2.2. Reintroductions

Understanding the dynamics of northern bristlebird grassy forest habitat is a vital part of pre-reintroduction habitat assessments. During this thesis I have shown that northern bristlebird persistence has been strongly influenced by the extent of habitat available, the condition of the grassy understorey present and the fire regime of a site. For reintroductions to be successful it will be important to reintroduce individuals to sites where these characteristics are present and can be maintained. Based on my findings, my recommendations for northern bristlebird reintroductions are:

 In Chapter 3 I found that northern bristlebird presence was highly spatially autocorrelated, with sites more likely to contain birds if they were close to other occupied sites. This has important implications for future reintroductions. Proximity to wild territories is likely to promote dispersal and breeding between captive and wild individuals, helping to improve genetic diversity across them. I therefore suggest priority release sites should be chosen that are close (within 1 km) to currently occupied territories.  Northern bristlebird presence was strongly associated with large habitat patches (Chapter 3), and I recommend that reintroductions should focus on sites where high quality grassy habitat is at least 40 ha in extent. Smaller sites could be considered if they have high quality grassy habitat and are close to other highly quality patches: if they are not, they are unlikely to be viable, with high direct mortality risk from fire events, and low connectivity with other suitable habitat patches for dispersal and interbreeding.  Sites with tall, thick grass cover, with a high occurrence of large tussocks should be prioritised for reintroductions over large patches that do not have appropriate grassy condition.  Northern bristlebird habitat has declined significantly since the 1980s and the subsequent isolation of suitable habitat means that even following reintroduction actions subpopulations will be highly fragmented across our study region. For northern bristlebird to persist, attempts need to be made to increase the extent of, and connectivity between remnant high quality grassy patches. Without connectivity between northern bristlebird subpopulations, limited dispersal and gene flow will occur.

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6.3. Study limitations

Field constraints meant that not all historically occupied sites could be sampled for this thesis. As a result, I only surveyed known historical and current northern bristlebird locations that were still within eucalypt forest that had some grassy elements in the understorey. Some historical sites were left out of the study because they have totally reverted to rainforest. Because the goals of this thesis were highly management orientated, I avoided focussing on sites that were clearly no longer suitable for northern bristlebird reintroduction. This decision meant that I was limited in sites that had long periods of fire absence, which may have reduced the strength of interactions between fire and habitat characteristics modelled in Chapter 5. Field constraints also meant that grassy habitat patches that had no known historical records of northern bristlebird were not sampled. Inclusion of such sites in my sampling design may have provided more detailed information on specific habitat characteristics that are associated with northern bristlebird presence, as current patterns of persistence may be more influenced by threat distribution and refuge habitat rather than habitat suitability (Osborne & Seddon, 2012).

A major constraint on testing whether food limitations are occurring within northern bristlebird habitat arises from not being able to study the diet and actual invertebrate species consumed by the birds. I was not able to study the diet of the northern bristlebird because as a cryptic ground-dwelling bird that lives in a thick grassy habitat it is very hard to see. In addition the very small remaining numbers of northern bristlebirds in the wild, together with the sensitivity of nesting birds to disturbance, made observations of foraging and prey preferences practically and ethically unfeasible. Because of this, I was limited to using availability of invertebrates in the habitat as a proxy for food quality. This measure does not necessarily represent what is actually consumed by northern bristlebirds, so caution is recommended in the interpretation of potential dietary limitation. In addition, the type of invertebrate sampling that I used, limits measure of food availability to the ‘standing crop’, i.e. the amount of food availability after prey items have already been consumed by birds living at a site (Fretz & Escalante, 2002; Hutto, 1990). This further restricts interpretations of food availability, as we may only be sampling the remaining, less preferred food resources (Hutto, 1990).

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Chapter 4 relied on evidence from southern bristlebird research that invertebrates were the major component of eastern bristlebird diet (Gibson & Baker, 2004). Considering how different northern bristlebird habitat is with a dominant grassy understorey, seeds may play a more important role in their diet, particularly during periods of very hot or very cold weather, or times of environmental stress when invertebrate availability is affected. As such, there may be untested dietary limitations occurring within northern bristlebird habitat. If seed is more important to northern bristlebirds, the loss of tussocks identified in Chapter 5 may be contributing to a food limitation for northern bristlebirds.

Invertebrates are highly variable in abundance, and they respond readily to environmental conditions (Kremen et al., 1993). Changes in climatic conditions means assessing community composition can be challenging, and requires long-term datasets to fully determine invertebrate availability in ecosystems. This high natural variability in invertebrates is likely to have limited my ability to fully assess invertebrate assemblages in northern bristlebird habitat. Further monitoring of insects, particularly between breeding and non-breeding territories, which is difficult given the small number of breeding sites confirmed, or during the summer and winter seasons which weren’t sampled, may shed further light on invertebrate resource patterns and potential limitations to northern bristlebirds.

6.4. Future research priorities

While the research presented in this thesis has advanced our knowledge on northern bristlebird persistence and habitat dynamics, several questions still remain that need to be addressed for future reintroductions and recovery to be successful. In Chapter 5 and the updated habitat model (Fig 6.1), I identified that long periods of fire absence can decrease tussock cover within northern bristlebird habitat. Tussocks are an important breeding resource (Holmes, 1989; Rohweder, 2000), and with the reduced extent of tall thick grassy habitat (Chapter 5), restoring tussock areas will be an important part of habitat restoration. Identifying tussock seed bank properties and methods for regenerating healthy tussocks within the grassy understorey would be a useful endeavour for future research.

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The ability of the northern bristlebird to use and disperse through rainforest following fire events is unknown. Research on this may provide useful insights into which habitat features promote individual survival or recolonization of burnt areas. Data on microhabitat selection by birds in different seasons and during the breeding period could also allow more accurate assessment of limiting resources. For example, radio-tracking of southern bristlebirds in Jervis Bay provided important information on habitat selection preferences (Baker, 2009), movement patterns (Bain et al., 2012) and population responses to translocations (Bain & French, 2009). Future reintroductions of northern bristlebirds would provide the perfect opportunity to use radio-tracking to investigate how northern bristlebirds use the landscape in response to fluctuating resource availability and fire.

This thesis has largely focussed on collecting ecological data to inform managers on actions that are most likely to maintain and improve habitat condition and northern bristlebird conservation. An important next step in the conservation of the bird would be a detailed prioritisation of potential release sites (Laws & Kesler, 2012) using the findings and data presented in this thesis. Release sites need to be carefully considered in terms of their suitability for northern bristlebirds and their ability to be managed effectively. Throughout this thesis I have provided relevant information for such a prioritisation.

6.5. Conclusion

The northern bristlebird has declined significantly in the last 30 years and will now require intensive intervention to persist. Despite this, there has previously been little scientific research on the habitat and fire characteristics needed to reverse the decline and promote long term persistence. Based on my findings, I have provided a range of population- specific recommendations for habitat management and future northern bristlebird reintroductions. From this information, I believe we now have a good understanding of northern bristlebird habitat requirements and how to maintain these using prescribed fire. The next step is to implement these management actions, and supplement the existing wild population with reintroductions from the captive breeding programme. Reintroduction of northern bristlebirds will be vital for survival of the population, as even with successful

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habitat rehabilitation, this current population size is almost certainly too small to be viable without supplementation.

As a result of the research presented in this thesis, along with the ongoing efforts and expertise of a dedicated recovery team, captive breeding program, and passionate land managers, we have the tools and knowledge to save the northern bristlebird. It is now just a matter of getting on with it.

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Appendix

Location First Historic Holmes Sites Surveyed recorded Territories Territories Conondale Range 1984 3-4 3 2 2 Grandchester* 1965 1 0 0 Cunninghams Gap 1952 23 0 4 2 Spicers Gap 1958 >6 35 11 8 Mt Maroon¹ 1977 1 0 1 0 Mt Barney¹ 1955 1 3 1 0 Pinnacle 1985 1 1 1 0 Surprise Rock¹ 1943 6-12 0 3 0 Sarabah Range 1930 4 1 3 3 Stretchers Track 1989 7 3 2 Mt Gipps/Richmond Gap 1982 13 31 13 14 Bald knob² 1982 1 1 1 0 Long Creek 1988 1 3 2 1 Green Pigeon² 1982 2 2 1 0 Serong Spur 1984 4 2 2 1 Mt Hutley 1984 4 7 4 3 Grassy Spur 1984 2 4 3 2 Cougal² 1984 1 3 2 0 Razorback Mt 1982 2 0 1 0 Mt Burrell 1982 1 0 1 0 Big Scrub/Lismore* 1866 0 0 Dorrigo Plateau* 1893 0 0 Wootton* 1922 0 0 TOTAL 103 59 38 ¹ Heath vegetation ² Completely rainforest, no grassy forest left at location * Presumed long extinction by Holmes (1989)

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Survey report Birds Population Territories Year Survey effort reference recorded estimate recorded 1982 Holmes (1982) 95 115 Unknown 1988 Holmes (1989) 154 154 103 115 sites Lamb (1993); Hartley 1992 24 31 88 sites (81 original + 7 new) & Kikkawa (1992) 1996 Holmes (1997) 32 36 30 Stewart (1997); 1997 23 14 117 sites (103 original + 14 new) Stewart (1998) 1999 Rowheder (1999) 40 47 25 2000 Sandpiper (2000) 37 29 24 118 sites (62 original + 56 new) 2001 Sandpiper (2001) 22 30 17 77 sites (57 original + 20 new) 2001 Holmes (2001) 17 13 2002 Sandpiper (2003) 14 19 16 65 sites (51 original + 14 new) 2003 Sandpiper (2005) 12 8 2005 Sandpiper (2005) 13 15 7 132 sites (77 original + 55 new) 2006 Sandpiper (2006) 18 28 24 122 sites (56 original + 66 new) 2007 Sandpiper (2008) 14 16 12 117 sites (40 original + 77 new) 2010 Sandpiper (2010) 10 22 11 67 sites (25 original + 42 new) 2011 Sandpiper (2012) 13 13-35 7 54 sites (24 original + 30 new) 2014 Wildsearch (2014) 13 13-30 7 81 sites (29 original + 52 new) 2016 Wildsearch (2016) 18 38 9 62 sites (29 original + 33 new)

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Appendix 4. Fire history information used to model the influence of fire on northern bristlebird habitat characteristics at current and historically occupied sites.

Region Site Bristlebird Time since Total Recent fire Mean fire Longest fire All fires presence fire (yrs) number intervals interval interval (yrs) of fires (yrs) (yrs) Border Ranges BR_BK 0 14 1 12 11.67 20 2002; 2000 Border Ranges BR_CP1 1 1 6 5.67 6 14 2015; 2001; 1999; 1996; 1986; 1980 Border Ranges BR_CP2 0 5 3 6.33 8.75 17 2011; 2001; 1997 Border Ranges BR_D 1 11 4 8.67 6.5 8 2005; 2001; 1993; 1990 Border Ranges BR_F 0 31 1 31 31 31 1985 Border Ranges BR_G2 1 1 3 3.67 5.33 11 2005; 2001; 1990 Border Ranges BR_G3 1 1 2 8 3 20 2005; 2000 Border Ranges BR_G4 0 16 3 10.33 10.33 16 2000; 1990; 1985 Border Ranges BR_GS1 1 5 3 11.67 9 9 2011; 2002; 1981 Border Ranges BR_GS2 1 5 3 11.67 9 9 2011; 2002; 1981 Border Ranges BR_H1 1 3 6 8.67 5.83 13 2013; 2000; 1990;1987; 1984; 1981 Border Ranges BR_H2 0 3 6 8.67 5 13 2013; 2000; 1990; 1987; 1985; 1984; 1981 2011; 2000; 2006; 2002; 2000; 1998; 1991; Border Ranges BR_HW 1 5 7 5.33 3.38 7 1987; 1985 Border Ranges BR_L 0 7 2 23.5 23.5 36 2009; 1969 Border Ranges BR_M 1 0 5 7.67 8 14 2016; 2007; 2000; 1986; 1984 Border Ranges BR_M1 1 0 4 3 7.5 21 2016; 2015; 2007; 1986 Border Ranges BR_R1 1 2 4 5 7.75 16 2014; 2009; 2001; 1985 Border Ranges BR_R2 1 2 5 5 6.2 14 2014; 2009; 2001; 1987; 1985 Border Ranges BR_R3 0 2 6 5 5.83 14 2014; 2009; 2001; 1987; 1985; 1981 Border Ranges BR_R4 1 2 3 7.33 7.33 14 2014; 2002; 1994 Border Ranges BR_R5 0 1 3 7 13 22 2015; 2002 Border Ranges BR_SC 0 3 5 8.33 7 13 2013; 2000; 1991; 1983; 1981 Border Ranges BR_SS1 1 0 3 3.33 5 26 2016; 2012; 2006 Border Ranges BR_ST1 0 0 4 8.67 1.93 7 2016; 2009; 2004; 2002; 2001 Border Ranges BR_ST2 0 0 4 5 1.93 7 2016; 2009; 2004; 2002; 2001

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A4 cont. Region Site Bristlebird Time since Total Recent fire Mean fire Longest fire All fires presence fire (yrs) number intervals interval interval (yrs) of fires (yrs) (yrs) 2011; 2008?; 2005; 2001; 2000; 1993; Border Ranges MR1 1 5 8 5 4 7 1990; 1984 2015; 2012; 2008; 2007; 2006; 2001; 1999; Conondales C1 0 1 15 2.67 2.27 5 1996; 1995; 1992; 1990; 1986; 1985; 1984; 1982 2015; 2007; 2006; 2003; 1996; 1990; 1987; Conondales C2 0 1 10 3.33 3.4 8 1986; 1985; 1982 Green Pigeon GP 0 31 2 17.5 17.5 31 1985; 1981 Helmut H 1 9 1 7 9 15 2013; 2007 Lamington L_BB 0 36 0 36 36 36 pre 1980s 2010; 2007; 2006; 2001; 1997; 1990; 1989; Lamington L_DC1 0 6 8 3.33 4 7 1984 Lamington L_DC2 0 6 4 8.67 7.75 17 2010; 2007; 1990; 1985 Lamington L_S 0 0 5 6.67 8.4 14 2016; 2007; 2004; 1990; 1982 Main Range MR_Ma 0 2 6 4.67 6 11 2014; 2006; 2002; 1991; 1985; 1980 Main Range MR_Mi 0 2 5 8.33 7.2 12 2014; 2003; 1991; 1986; 1980 Main Range MR_S1 0 0 5 12.67 9.8 14 2016; 2002; 1991; 1984; 1980 Main Range MR_S2 1 2 7 3.33 4.57 11 2014; 2011; 2006; 2002; 1991; 1985; 1984 Main Range MR_S3 0 10 3 9.67 9.67 15 2006; 2002; 1987 Main Range MR_S4 0 5 5 3.33 5 11 2011; 2009; 2006; 2002; 1991 Main Range MR_S4B 0 5 4 3.33 3.5 22 2011; 2007; 2006; 2002 Main Range MR_S5 0 10 2 12 7 22 2006; 2002 Main Range MR_S6 0 2 5 3.33 5 11 2014; 2011; 2006; 2002; 1991 Main Range MR_S8 1 0 2 16 16 32 2016; 1984 * Sites not surveyed for vegetation or invertebrates, excluded from analysis

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