Michael T. Lohr B.S

Response of Australian Boobooks (Ninox boobook) to threatening processes across urban, agricultural, and woodland ecosystems

Michael T. Lohr B.S. The Pennsylvania State University M.S. The University of Delaware

Thesis Submitted for the degree of Doctor of Philosophy in the School of Science Edith Cowan University

November 2019

“One of the penalties of an ecological education is that one lives alone in a world of wounds. Much of the damage inflicted on land is quite invisible to laymen. An ecologist must either harden his shell and make believe that the consequences of science are none of his business, or he must be the doctor who sees the marks of death in a community that believes itself well and does not want to be told otherwise.”

- Aldo Leopold, “A Sand County Almanac”

Abstract The effects of habitat fragmentation on native wildlife can vary depending on the type of land use occurring in the matrix between remaining habitat fragments. I used Australian boobooks (Ninox boobook) in Western Australia to investigate interactions between matrix type and four different potential threatening processes: secondary poisoning by anticoagulant rodenticides (ARs); limitation of juvenile dispersal and impacts on spatial genetic structure; breeding site availability; and infection by the parasite Toxoplasma gondii.

I also conducted a literature review on the use and regulation of ARs in Australia and published accounts of non-target impacts in order to contextualise exposure patterns observed in boobooks. The review revealed records of confirmed or suspected poisoning across 37 vertebrate species in Australia. World literature relating to AR exposure in reptiles suggests that they may be less susceptible to AR poisoning than birds and mammals. This relative resistance may create unevaluated risks for wildlife and humans in Australia where reptiles are more abundant than in cooler regions where AR exposure has been studied in greater depth.

I analysed AR residues in boobook livers across multiple habitat types. Second generation anticoagulant rodenticides were detected in 72.6% of individuals sampled. Total AR concentration correlated positively with the proportion of urban land use within an area approximately the size of a boobook’s home range centred on the point where the sample was collected. ARs originating in urban habitat probably pose a substantial threat to boobooks and other predatory wildlife species.

No spatial genetic structure was evident in boobooks across habitat types. I observed one individual dispersing at least 26km from its natal home range across urban habitat. The apparent permeability of anthropogenically altered landscapes probably explains the lack of spatial genetic structure and is likely related to the observed ability of boobooks to use resources in both urban and agricultural matrices.

Boobooks did not appear to be limited by the availability of suitable nesting sites in urban or agricultural landscapes. Occupancy did not change significantly over the duration of the study in remnants provided with artificial nest boxes in either landscape type.

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However, in one instance, boobooks successfully used a nest box located in an urban bushland. Nest boxes may be a useful management tool in highly-altered areas where natural hollows are unavailable.

Toxoplasma gondii seropositivity in boobooks did not vary significantly by landscape type but was more prevalent in individuals sampled during cooler wetter times of year. Risk of exposure due to greater cat abundance in urban and agricultural landscapes may be offset by creation of environmental conditions less favourable to the survival of T. gondii oocysts in soil.

Taken together, this body of research demonstrates variation in relationships between different types of habitat fragmentation and threatening processes related to fragmentation. This research also raises questions about how habitat fragmentation is discussed and studied in the context of species which are capable of making extensive use of matrix habitat. I recommend greater consideration of the concept of “usable space” when studying fragmentation impacts in habitat generalists.

Declaration

I certify that this thesis does not, to the best of my knowledge and belief:

i. incorporate without acknowledgment any material previously submitted for a degree or diploma in any institution of higher education; ii. contain any material previously published or written by another person except where due reference is made in the text; or iii. contain any defamatory material. iv. I also grant permission for the Library at Edith Cowan University to make duplicate copies of my thesis as required.

Michael T. Lohr 06/11/2019

Acknowledgments

I would first like to thank my supervisors Dr. Rob Davis and Dr. Allan Burbidge. Their insights into navigating the complex ecosystem that is conservation research in Western Australia are greatly appreciated. I sincerely appreciate the free rein they gave me in exploring a series of sometimes unconventional side projects. These opportunities have proved invaluable. Rob’s willingness to meet at length to discuss new opportunities and troubleshoot occasional difficulties made the entire PhD experience easier and more enjoyable.

Cheryl, your accommodation of my bizarre nocturnal field schedule, financial support, tolerance for endless monologues about anticoagulant rodenticides, and R code are what made this whole thing actually work. Thank you. I look forward to having our life back in the near future.

Many thanks to the large number of people and organisations willing to hold their collective noses and accumulate dead owls for me. This PhD would not have been possible without your efforts. I hope to continue to do my part to convert the smelly data you collected into meaningful conservation actions. Samples were contributed by Kanyana Wildlife Rehabilitation, Native Animal Rescue, Native ARC, Nature Conservation Margaret River Region, Eagles Heritage Wildlife Centre, and many individual volunteers especially Steve Castan, Simon Cherriman, Angela Febey, Warren Goodwin, Amanda Payne, Stuart Payne, and Boyd Wykes.

Many people provided help on long nights of owl surveys and nest box checks including: Casper Avenant, Rachele Bernasconi, Jakeb Cumming, Angela Febey, Sian Glazier, Melissa Hetherington, Tyson Isles, Michael Just, Candice Le Roux, Gabe Mach, Paul Radley, Calan Rance, Geoffrey Schoonakker, Nakisa Shahrestani, Lia Smith, Steven Spragg, Paula Strickland, and Mitch Wright.

I am particularly grateful to Simon Cherriman, whose enthusiastic assistance in preliminary field work helped me build confidence in working with these amazing birds. His subsequent nest box design, construction, and installation and advice on interpretation were critical to the nest box chapter. The inclusion of Simon’s photo in the title page of this

vi dissertation is a testament to the quality of his photography and the mileage I have gotten out of his photos of my work. I sincerely hope I can repay my debt as he continues his PhD and I look forward to future and ongoing collaborations.

I wish to express my sincere thanks to Dr. Jamie Tedeschi for her patience and expertise in introducing me to the world of genetic analysis and to Louise Pallant and A/Prof. Annette Koenders for their advice and assistance on serological testing. Training a field ecologist to do lab work is surely a painful experience and I am grateful that they attempted it.

I particularly appreciate Jerry Olsen contributing data from his boobook banding projects as well as helpful advice and friendly correspondence throughout my PhD.

I also thank Ben Jones and Yvonne Sitko for helping me to communicate the results of my work to the public. Without their efforts, much of my work would not have made it to the people who can actually use it.

I especially thank Rachele Bernasconi, Casper Avenant, Melissa Karlinski, Emily Lette, Rosh McCallum, and Charlie Phelps for their moral support and for tolerating my eccentricity, frightening desktop, and questionable musical taste through the writing process. Your contributions to my sanity were critical to getting this dissertation finished.

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Statement of contribution of others

Research Funding The Holsworth Wildlife Research Endowment via The Ecological Society of Australia BirdLife Australia Stuart Leslie Bird Research Award Edith Cowan University School of Science Postgraduate Student Support Award Eastern Metropolitan Regional Council’s Healthy Wildlife Healthy Lives program The Society for the Preservation of Raptors Sian Mawson Stipend Edith Cowan University Postgraduate Research Scholarship Edith Cowan University Merit Award Supervision Dr. Robert A. Davis Dr. Allan H. Burbidge Field Assistance Casper Avenant, Rachele Bernasconi, Simon Cherriman, Jakeb Cumming, Angela Febey, Sian Glazier, Melissa Hetherington, Tyson Isles, Michael Just, Candice Le Roux, Gabe Mach, Paul Radley, Calan Rance, Geoffrey Schoonakker, Nakisa Shahrestani, Lia Smith, Steven Spragg, Paula Strickland, Mitch Wright Laboratory Technical Assistance and Advice A/Prof. Annette Koenders, Louise Pallant, Dr. Jamie Tedeschi, Co-Authors Dr. Janet Anthony, Dr. Allan H. Burbidge, Simon Cherriman, Dr. Robert A. Davis, Dr. Siegfried Krauss, Dr. Cheryl A. Lohr, A/Prof. Peter B. S. Spencer The research included in this dissertation is my original work. I conceived and developed all hypotheses, led all field work, designed or conducted the majority of analysis,wrote all first drafts, and made the majority of edits to subsequent drafts. The co- authors listed above contributed to one or more chapters in at least one of the following ways: advice on experimental design, assistance in fieldwork, data analysis, and editing of drafts. I am the lead author on all published articles and manuscripts. My roles in each chapter are detailed in the “Co-author Statements” section.

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Publications arising from this research

I am submitting this thesis as a thesis by publication. Chapters 2 and 3 are reformatted versions of the published journal articles. A single reference list is provided for the entire thesis following the final chapter. No permission is needed to reproduce these articles as part of a PhD thesis. The first pages of published chapters 2 and 3 can be found in the section entitled “Copies of original publications”

Chapter 2

Lohr, M. T., and R. A. Davis (2018). Anticoagulant rodenticide use, non-target impacts and regulation: A case study from Australia. Science of the Total Environment. 634:1372– 1384.

Chapter 3

Lohr, M. T. (2018). Anticoagulant rodenticide exposure in an Australian predatory bird increases with proximity to developed habitat. Science of the Total Environment. 643:134–144.

Chapter 4

Lohr, M. T., P. B. S. Spencer, S. Krauss, J. Anthony, A. H. Burbidge, and R. A. Davis. Widespread genetic connectivity in Australia’s most common owl, despite extensive habitat fragmentation. The Condor: Ornithological Applications. (In Preparation).

Chapter 5

Lohr, M. T., S. Cherriman, A. H. Burbidge, and R. A. Davis. Artificial nest box supplementation does not affect Australian boobook (Ninox boobook) occupancy in fragmented habitats in south-western Australia. Wildlife Research. (In Review).

Chapter 6

Lohr, M. T., C. A. Lohr, A. H. Burbidge, and R. A. Davis. Toxoplasma gondii seropositivity across urban and agicultural landscapes in an Australian owl. Veterinary Parasitology. (In Preparation).

Table of Contents

Abstract ...... iii Declaration ...... v Acknowledgments ...... vi Statement of contribution of others ...... viii Publications arising from this research ...... ix Table of Contents ...... x List of Figures ...... xiv List of Tables ...... xvi A note on nomenclature ...... xvii Chapter 1 Introduction ...... 1 Chapter 2 Anticoagulant rodenticide use, non-target impacts and regulation: A case study from Australia ...... 10 Abstract ...... 10 Introduction ...... 11 Aims ...... 12 Methods ...... 12 Results and Discussion ...... 13 Literature Survey ...... 13 Anticoagulant Exposure of Non-target Wildlife in Australia ...... 14 Current Uses in Australia ...... 25 Unique Considerations in Australia...... 30 Conclusions and Recommendations ...... 38 Acknowledgements ...... 40 Appendix 2.A. Definitions of Schedules applying to all Anticoagulant Rodenticides Registered in Australia from (Australian Government Department of Health: Therapeutic Goods Administration, 2017) ...... 41 Chapter 3 Anticoagulant rodenticide exposure in an Australian predatory bird increases with proximity to developed habitat ...... 42 Abstract ...... 42 Introduction ...... 42 Methods ...... 44 Specimen Collection ...... 45 Rodenticide Analysis ...... 45 Statistical Analysis ...... 46

Exposure Thresholds ...... 47 Spatial Analysis ...... 48 Results ...... 49 Discussion...... 57 Individual Rodenticides ...... 58 Rodenticide Thresholds ...... 61 Spatial Correlations ...... 62 Seasonal Differences ...... 66 Rodenticide in fledglings ...... 67 Conclusion ...... 69 Acknowledgements ...... 69 Chapter 4 Widespread genetic connectivity in Australia’s most common owl, despite extensive habitat fragmentation ...... 71 Abstract ...... 71 Introduction ...... 72 Habitat Fragmentation, Connectivity, and Genetic Structure ...... 72 Genetic Responses of Predatory Birds to Fragmentation...... 72 Declines in Australian Boobook Abundance ...... 73 Boobook Movement and Responses to Fragmentation ...... 74 Methods ...... 76 Juvenile Dispersal ...... 76 Genetic Sample Collection ...... 76 Genetic Analysis ...... 78 Statistical Analysis ...... 80 Results ...... 81 Direct Measurement of Dispersal ...... 81 Indirect Estimation of Dispersal ...... 84 Discussion...... 88 Acknowledgments ...... 91 Appendix 4.A A complete listing of the samples used in the analysis of microsatellite DNA polymorphisms, including the identification number (Individual ID), sample source, collection dates, collection locations (decimal lat/long), sampling locations/regions and age at sampling of Australian Boobooks used in this study. HY=hatch year, SY=second year, AHY=after hatch year, ASY=after second year...... 93 Appendix 4.B CLUMPAK results showing median values of the natural log of the probability of the number of genetic clusters (K=1-6) in Australian Boobooks sampled in Western Australia...... 99 Appendix 4.C STRUCTURE HARVESTER output indicating the highest probability for K=1 in boobooks sampled in Western Australia...... 99

Chapter 5 Artificial nest box supplementation does not affect Australian boobook (Ninox boobook) occupancy in fragmented habitats in south-western Australia ...... 100 Abstract ...... 100 Introduction ...... 101 Nest Competition and Predation ...... 102 Impacts of Nest Boxes in Conservation ...... 103 Knowledge Gaps...... 104 Methods ...... 106 Study Sites ...... 106 Surveys ...... 107 Nest box construction and placement ...... 108 Nest Box Monitoring ...... 111 Statistical Analysis ...... 112 Results ...... 112 Discussion...... 114 Surveys ...... 114 Nest Box Use ...... 115 Conclusion ...... 119 Acknowledgments ...... 119 Chapter 6 Toxoplasma gondii seropositivity across urban and agricultural landscapes in an Australian owl ...... 120 Abstract ...... 120 Introduction ...... 121 Effects of Toxoplasma gondii on Humans and Wildlife ...... 122 Predatory Birds and Toxoplasma gondii Infection ...... 123 Aims...... 124 Methods ...... 125 Sample Collection ...... 125 Serological Testing ...... 126 Statistical Analysis ...... 127 Results ...... 128 Discussion...... 131 Landscape Type ...... 133 Age ...... 133 Injury Status ...... 134 Season ...... 134

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Anticoagulant Rodenticide Exposure ...... 135 Acknowledgments ...... 136 Chapter 7 Summary, Synthesis, and Management Implications ...... 137 Summary of major findings: ...... 137 Objective 1. Critically review literature on anticoagulant rodenticide exposure in native wildlife in Australia to clarify its role as a threatening process...... 137 Objective 2. Investigate the relationship between exposure to anticoagulant rodenticides and urban and agricultural fragmentation...... 138 Objective 3. Determine if urban and agricultural fragmentation influence boobook genetic structure...... 138 Objective 4. Examine whether nest box supplementation increases site occupancy at unoccupied sites and whether this effect differs between urban and agricultural landscapes. 139 Objective 5. Explore patterns of Toxoplasma gondii seropositivity in boobooks across the urban, agricultural, and natural landscapes...... 139 Synthesis ...... 140 Management Recommendations ...... 144 Anticoagulant Rodenticides ...... 144 Nest Box Supplementation ...... 145 References ...... 146 Co-author Statements ...... 185 Copies of original publications ...... 189

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List of Figures Figure 3.1 Percentages of Southern Boobooks (n=73) in Western Australia exposed to rodenticides stratified by total rodenticide liver concentration (mg/kg) thresholds indicating potential outcomes...... 54 Figure 3.2 Percentages of Southern Boobooks (n = 73) exposed to multiple anticoagulant rodenticides in Western Australia...... 55 Figure 3.3 Mean total anticoagulant rodenticide concentration (mg/kg) in liver tissue of Southern Boobooks (n= 71) in Western Australia by season...... 56 Figure 4.1 Sample locations of genotyped Australian Boobooks (Ninox boobook) in Western Australia. (“metro” = urban and suburban areas of Perth represented by squares, “rural” = forested area surrounding the Perth Metropolitan area represented by an “x” , “Southwest WA” = forested areas to the south of Perth represented by triangles, “Wheatbelt” = highly-fragmented agricultural landscapes represented by crosses, and “other” = Goldfields and Pilbara regions, represented by black circles, ‘other’ = Goldfields and Pilbara regions of Western Australia)...... 77 Figure 4.2 A corellogram showing genetic correlation values (r) as a function of distance (kms) using eight microsatellite markers in a subset of Australian Boobooks (Ninox boobook) n=98 from the Perth metropolitan area, adjacent exurban areas and the Perth Hills. U and L are 95% confidence intervals around the null hypothesis of no spatial genetic structure. No significant genetic structure is shown at any distance class...... 83 Figure 4.3 Principal coordinate analysis results based on eight microsatellite loci in Australian Boobooks (Ninox boobook) in Western Australia. Clustering does not correspond to potential populations and is driven by two common alleles and their heterozygotes at the locus Nst15. Blue = 161/161, Green = 161/uncommon allele, Purple = 163/161, Orange = 163/uncommon allele, Red = 163/163, Black = no result...... 83 Figure 4.4 Principal coordinate analysis results based on seven microsatellite loci (i.e. no Nst15 – see Fig 3) in Australian Boobooks in Western Australia. No clustering is apparent across or within six sampled regions (“Exurbs” = areas immediately surrounding but not within the Perth Metropolitan area, “Perth Hills” = an area of continuous forest east of Perth, “Perth Metro” = urban and suburban areas of Perth, ‘Remote WA’ = Goldfields and Pilbara regions of Western Australia, “Southwest WA” = forested areas to the south of Perth, “Wheatbelt” = highly-fragmented agricultural landscapes existing primarily between the “Remote” region and all other regions)...... 85 Figure 4.5 Visualization of Australian Boobooks (Ninox boobook) sampled from six regions in Western Australia (“Exurbs” = areas immediately surrounding but not within the Perth Metropolitan area, “Perth Hills” = an area of continuous forest east of Perth, “Perth Metro” = urban and suburban areas of Perth, ‘Remote WA’ = Goldfields and Pilbara regions of Western Australia, “Southwest WA” = forested areas to the south of Perth, “Wheatbelt” = highly-fragmented agricultural landscapes existing primarily between the “Remote” region and all other regions) using the STRUCTURE results from CLUMPAK comparing number of inferred genetic clusters (K) from 1-6. The data support a single genetic cluster. Each line represents an individual. The proportion of colours in each line represents the proportion of membership of each individual in each cluster...... 86 Figure 4.6 Plot of Evanno et al.’s (2005) delta K (ΔK) based on inferred genetic clusters (populations) ranging from 2 to 5 in Australian Boobooks (Ninox boobook) sampled from Western Australia...... 87

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Figure 5.1 Locations of survey sites in in southwestern Western Australia: urban landscapes in the Perth Metropolitan Area, continuous bushland in the Perth Hills, and agricultural landscapes within a 60km radius of Kellerberrin, Western Australia...... 107 Figure 5.2 Attachment system used to hang nest boxes used in this study...... 110 Figure 5.3 A nest box installed in one of the remnant bushlands in an agricultural landscape in Western Australia...... 111 Figure 6.1 Seasonal Toxoplasma gondii seroprevalence in Australian Boobooks (Ninox boobook) in Western Australia. Width of the bars is representative of sample size...... 130 Figure 6.2 Toxoplasma gondii seroprevalence in meat juice from deceased Australian Boobooks (Ninox boobook) in Western Australia in four different categories of anticoagulant rodenticide exposure (A= ≤ 0.01 mg/kg, B=0.01 mg/kg – 0.10 mg/kg, C 0.10 mg/kg - 0.50mg/kg, D ≥ 0.50mg/kg) Width of the bars is representative of sample size...... 131

List of Tables Table 2.1 Numbers and categories of publications relating to anticoagulant rodenticides in Australia...... 14 Table 2.2 Accounts of non-target AR toxicity in Australian wildlife. *Authors do not specify how poisoning was verified ...... 16 Table 2.3 Anticoagulants currently approved for vertebrate pest control in Australia. Some anticoagulants are assigned different schedules dependant on formulation. *Some disagreement exists as to whether these should be treated as first or second generation anticoagulants †Warfarin is used therapeutically in humans as a blood thinner...... 23 Table 3.1 Limit of detection (LOD), limit of quantification (LOQ), average recovery, and relative standard deviation (RSD) for eight ARs in a spiked chicken liver matrix...... 46 Table 3.2 Percentage exposure, mean exposure and total detection of eight different anticoagulant rodenticides in livers of 73 Southern Boobooks in Western Australia...... 50 Table 3.3 Published rates of multiple second generation anticoagulant rodenticide exposure and percentages of individuals with exposure above two thresholds in predatory birds...... 51 Table 3.4 Akaike information criterion (AIC) ranking of models of the association between percentage of single land use types within buffers around collection points and total anticoagulant rodenticide liver concentration in Southern Boobooks (n= 66) in Western Australia at three different spatial scales (Big=2827.4 ha buffer, Mid=145.1 ha buffer, Small=7.3 ha buffer...... 57 Table 4.1 The characteristics of the primers from 15 microsatellite loci amplified in Australian Boobooks (Ninox boobook) from Western Australia using primers adapted from (Hogan et al. 2007, 2009)...... 79 Table 4.2 Records of date a bird was tagged, its location, days and distances elapsed between capture and recovery of Australian Boobooks (Ninox boobook) banded as fledglings in Australia. Data from the Australian Capital Territory (ACT) and Queensland sourced from the Australian Bird and Bat Banding Scheme (http://www.environment.gov.au/science/bird-and-bat-banding). Western Australian data from re-sightings and recoveries of boobooks captured as part of this study...... 82 Table 4.3 Analysis of Molecular Variance (AMOVA) results using six regional groups of Australian Boobooks (Ninox boobook) in Western Australia as populations...... 82 Table 4.4 Genetic diversity parameters for Australian Boobooks (Ninox boobook) in six regions in Western Australia derived from eight microsatellite loci. Mean number of genotyped individuals (N), mean number of alleles per locus (NA), mean number of effective alleles (NE), mean observed heterozygosity (HO), mean unbiased expected heterozygosity (uHE)...... 87 Table 4.5 Pairwise Fst and estimated number of migrants per generation (NM) between all geographic regions of Australian Boobooks (Ninox boobook) sampled in Western Australia...... 87 Table 4.6 Pairwise estimates of Jost's DST (below diagonal) and associated P values (above diagonal) for Australian Boobooks (Ninox boobook) sampled in five regions of Western Australia...... 88 Table 5.1 Annual change in occupancy of Australian Boobooks at continuous bushland sites and sites with and without supplemental nest boxes in remnant woodland in urban and agricultural landscapes in Western Australia...... 113 Table 5.2 Number of nest boxes used by bird species in urban and agricultural remnant woodlands across two years in Western Australia...... 114 Table 6.1 Factors associated with Toxoplasma gondii seroprevalence in Australian Boobooks (Ninox boobook) in Western Australia...... 128

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A note on nomenclature Over the course of my PhD, there have been several changes in the accepted common name and scientific name of the species I have focused on. Previous literature frequently referred to the species as the Southern Boobook (Ninox novaseelandiae). However, more recently, authors have recognised a split between individuals on the Australian mainland and those in New Zealand and Tasmania (Olsen, 2011a). Subsequent simultaneous examination of genetic and bioacoustics evidence supports this split (Gwee et al., 2017). I accept this evidence and use Ninox boobook throughout the thesis to describe the birds that I studied. Following other splits suggested by Gwee et al. (2017) the International Ornithological Congress changed the common name “Southern Boobook” to “Australian Boobook” on January 20, 2019. This was done to distinguish boobooks found on the Australian mainland from other newly recognised species in the Lesser Sunda Islands. Accordingly, I have used “Australian Boobook” throughout my dissertation except in chapters 2 and 3 which were published prior to this change. In these chapters I have retained the old common name “Southern Boobook” to maintain consistency between my dissertation and the published journal articles.

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Chapter 1 Introduction

The availability of resources necessary for survival is the key factor driving spatial distribution and diversity in wildlife species (e.g. Isaac et al., 2014b). In highly human- altered landscapes, these resources are often restricted to remnant patches of native habitat and the value of these patches to native biodiversity is dependent on the continued ability of these patches to provide the resources required by native species (Harper et al., 2005). Fragmentation of areas of continuous natural vegetation by highly-altered landscapes impacts native wildlife through a variety of mechanisms including habitat loss (Ewers and Didham, 2006), isolation of remaining patches (Saunders et al., 1991), and degradation of remaining patches through edge effects (Collinge, 1996). The latter two mechanisms are strongly influenced by the types of land use that replace native vegetation on a landscape scale.

The process of habitat loss occurs through reduction in landscape-level availability of critical resources by conversion of what was previously relatively-undisturbed native habitat into another land use type (Collinge, 1996). Habitat loss occurs as part of the process of habitat fragmentation but the two phenomena have been conflated in some research, yielding overstated conclusions about the impacts of fragmentation (Haila, 2002). When viewed separately, models suggest that habitat loss has a substantially larger impact on the probability of species persistence than differences in spatial configuration of remaining patches (Fahrig, 1997). Rigorous investigations of fragmentation impacts must consider the interplay between fragmentation and habitat loss in order to distinguish which of the related processes is responsible for the effects under observation.

The influence of matrix attributes on landscape-level connectivity and dispersal of organisms between habitat patches was not immediately recognized in fragmentation ecology. Early investigations of the impacts of fragmentation on wildlife focused heavily on spatial distribution and size of habitat patches and largely adopted the paradigm of habitat fragments as islands, derived from MacArthur et al.'s (1967) theory of island biogeography (Haila, 2002). Accordingly, length of isolation, distance from other patches, and patch size were posited as the driving forces behind declines in biodiversity in habitat fragments

(Saunders et al. 1991) . While these factors do have some influence on species persistence within habitat patches, their repeated investigation has not led to substantive practical advances in conservation (Saunders et al. 1991). However, this line of investigation did contribute to the development of metapopulation theory (Levins, 1969) which provides a series of conceptual frameworks for understanding the persistence of populations with patchy distributions across a fragmented landscape based on patterns of connectivity between patches (Harrison and Taylor, 1997).

More recently, a number of studies have indicated that the type of matrix surrounding habitat fragments can significantly influence the ability of animals to disperse between patches, thus altering connectivity (Bender and Fahrig, 2005; Pither and Taylor, 1998; Ricketts, 2001) and potentially metapopulation dynamics. Watson et al. (2005) documented differing responses of woodland birds to habitat fragmentation in urban, peri- urban, and agricultural matrices. They did not explore the mechanisms causing these effects but suggested further research in this area. Further investigating the links between matrix type and landscape-level connectivity is a crucial component of integrating traditional views of landscape ecology with more current avenues of research.

A growing body of research indicates that the matrix between patches can have profound effects on species and communities within patches. Much of this impact occurs through degradation of key resources within remaining habitat patches (Hunter 2002). Mechanisms by which this degradation occurs include: increased abundance of disturbance- adapted alien species (Hansen and Clevenger, 2005); changes in light, temperature, and humidity (Matlack, 1993); and increased wind movement (Davies-Colley et al. 2000). These phenomena are especially pronounced in small and irregularly shaped patches which have larger areas of edge habitat in proportion to their interior areas (Collinge, 1996). In a review of the impacts of fragmentation, Saunders et al. (1991) argued that the attributes of the surrounding matrix have a greater influence on species persistence within patches than biogeographic factors and subsequent studies have elaborated on the important role that the surrounding matrix has on dynamics within fragments (Jules & Shahani 2003; Ewers & Didham 2006; Williams et al. 2006). More recent applied research has found that altering management practices within land use categories typically viewed as “hostile matrix” (e.g. suburban housing developments) can increase use of those habitats by native wildlife at

2 multiple trophic levels (Burghardt et al. 2009) reducing the “hostility” of the matrix to the point that it is usable habitat for at least some species. However, this work did not examine impacts within adjacent remaining habitat patches. Further exploration of within-patch impacts associated with specific matrix types is necessary for developing conservation strategies that operate effectively on a landscape scale.

The spatial configuration of land cover types in southwestern Western Australia is ideally suited to examining the effects of matrix type on the threatening processes associated with fragmentation. The Perth metropolitan area is separated from a large area of agricultural land use – frequently referred to as the wheatbelt – by a continuous band of largely-intact native woodland. Both urban and agricultural areas in the study area contain numerous patches of remnant native vegetation and exist on the same latitudinal gradient. In combination, these factors make this section of southwestern Western Australia an excellent candidate for this and future fragmentation studies examining the impacts of matrix type on wildlife utilizing habitat remnants. These features of southwestern Western Australia have facilitated a number of previous studies relating to the impacts of fragmentation on native wildlife in both urban (Davis and Wilcox, 2013; How and Dell, 1994; Krawiec et al., 2015) and agricultural (Hobbs and Saunders, 1991; Saunders, 1989; Saunders et al., 2014) areas.

Targeted examination of the impacts of fragmentation within taxonomic groups and feeding guilds is necessary because the effects of fragmentation can vary widely among different groups of organisms (Robinson et al., 1992). I selected a predatory species as a model because predators are more frequently extirpated as a result of fragmentation than animals at lower trophic levels as a result of their larger home range requirements and smaller population sizes (Didham et al., 1998; Duffy, 2003; Gilbert et al., 1998). Additionally, extirpation of predators can lead to trophic skew and resultant disruptions to food webs and ecosystem function. Removal of predators from an ecosystem can have impacts on biological systems that are as serious as much larger reductions in diversity of primary producers (Duffy, 2003). Among birds, predatory species have been observed to be at greater risk of extinction as a result of fragmentation (Leck 1979; Brash 1987; Carrete et al. 2009). As a consequence, it is crucial that we develop a better understanding of the threatening processes impacting predatory birds in urban areas and highly-modified

3 agricultural landscapes. While many threats to these species are understood qualitatively, few have been quantified in the field and, when they are, they are rarely addressed spatially on a landscape scale.

In recent years, important progress has been made in understanding how carnivorous birds are impacted by both agricultural and urban development. Responses of predatory birds to urban development are largely negative but can vary widely depending on the ecology of the species concerned and the type of modification to natural landscapes. For instance, Hager (2009) reviewed the literature on this topic and found many reports of impacts from electrocutions and collisions with anthropogenic objects and vehicles across a wide range of owl and raptor species. In some species, higher levels of mortality from vehicle collisions were associated with urban areas (Hager, 2009). Likewise, agricultural intensification has led to declines in carnivorous bird abundance resulting from loss of nesting sites, pesticide poisoning, and overgrazing of prey species habitat (Newton, 2004) as well as continental-scale decline across farmland bird species generally (Donald et al., 2001).

Conversely, a number of examples exist of generalist avian carnivores benefitting from urbanization and agricultural intensification. Eastern Screech Owls (Megascops asio) in Texas were found to exist at a higher density in a suburban area than in a rural area (Gehlbach, 1996). The suburban population also had higher adult survival, productivity, nest success, and stability than its rural counterpart (Gehlbach, 1996). Gehlbach (1996) attributed these differences to higher prey availability, increased climatic stability, and reduced numbers of avian predators in suburban areas relative to rural sites. Similarly, a study on Marsh Harriers (Circus aeruginosus) in Spain suggested that range-wide increases in marsh harrier abundance may be related to increased habitat suitability resulting from agricultural intensification (Cardador et al., 2011).

However, avian predator responses to urban and agricultural development can be complex and care must be taken when evaluating their impacts on a given species. Lesser kestrels in urban habitats in Spain suffered lower predation of adults and nestlings than their rural counterparts but nestlings in urban areas died of starvation more frequently (Tella et al., 1996). Similarly, Cooper’s Hawks (Accipiter cooperi) in urban areas of Tucson, Arizona occurred at higher densities, nested earlier, and had larger clutch sizes than their

4 exurban counterparts, likely as a result of high abundance of doves which made up the majority of their diet (Boal and Mannan, 1999). Despite occurring in high densities, nest success was significantly lower in urban areas, largely due to high rates of trichomoniasis in nestlings, and was not high enough to account for the stable or increasing number of adults observed in the urban area (Boal and Mannan, 1999). Consequently, urban Tucson appears to be an ecological trap for Cooper’s Hawks (Battin, 2004).

Within Australia, only a few studies have addressed landscape-level impacts of urban and agricultural development on predatory birds and most have focused specifically on Powerful owls (Ninox strenua). In Powerful Owls, one model suggested that high prey abundance in urban woodland fragments could create an ecological trap if prey availability serves as cue to preferentially establish territories in areas without adequate nesting hollows (Isaac et al., 2014a). This followed on from a previous model that predicted declining habitat suitability with urbanization in Powerful Owls (Isaac et al. 2013). One study of nightbird occurrence in southeastern Australia, found differing impacts of fragmentation on occurrence of several owl species (Kavanagh and Stanton, 2002). In this study, larger forest specialists were largely intolerant of fragmentation. Smaller generalist species, including boobooks, occurred over a wide range of fragmentation levels, but had lower occupancy rates in more fragmented habitats.

A few studies have attempted to examine the threatening processes impacting carnivorous birds in urban and agricultural landscapes in Australia and, again, most have focused on Powerful Owls. One study used shed feathers to examine the genetics of Powerful Owls and identified two instances of inbreeding in an area on the urban fringe of Melbourne (Hogan and Cooke, 2010). Cooke et al. (2006) found that food was not limiting Powerful Owl abundance along an urban to forest gradient. A study from an agricultural landscape noted a correlation between the use of a brodifacoum-based rodenticide in Queensland canefields and a decline in abundance of nesting pairs in seven owl species (Young and De Lai, 1997). However, several other threatening processes suspected to impact carnivorous birds remain largely unstudied in Australia and no studies, to my knowledge, have addressed the prevalence of these threatening processes across multiple types of anthropogenically altered habitat.

The sensitivity of rare species to threatening processes associated with development and their inherently small population size makes it difficult to directly and ethically study the impact and relative importance of these factors for those species. Australian Boobooks (Ninox boobook) provide an excellent model to quantify the spatial distribution of the threatening processes associated with fragmentation and highly altered landscapes. Boobooks are found in a variety of habitats in Australia and their basic biology and natural history is well documented. Of practical importance to this study, boobooks are common and widespread in the forests of southwest Western Australia (Liddelow et al. 2002). They also appear to be relatively resistant to some degree of fragmentation due to logging (Milledge et al. 1991; Kavanagh & Peake 1993; Kavanagh et al. 1995; Kavanagh & Stanton 2002) and may benefit from it in some cases (Kavanagh and Bamkin, 1995). Closely-related moreporks were detected at 80% of bushland patches in an urban area in NZ (Morgan and Styche, 2012). Trost et al. (2008) documented the use of a highly developed urban area as a winter home range by a female boobook. The documented use of native bushland, urban, and agricultural habitat types by boobooks allows their use as a model in investigating the influence of landscape type on the severity of the identified threatening processes.

However, reductions in boobook abundance have been observed or suspected following land clearing for agriculture (Leake, 1962; Masters and Milhinch, 1974; Saunders and Ingram, 1995) and urban development (Stranger, 2003). In one instance, the construction of a new road through five boobook territories led to the abandonment of three of the territories and enlargement of the remaining two (Olsen and Trost, 2007). Kavanagh & Stanton (2002) also observed lower occupancy rates in more fragmented habitats in southeastern Australia. Boobooks also appear to have undergone a significant range-wide decline between the first and second Atlas of Australian Birds (Barrett et al. 2003). BirdLife Australia’s (2015) “State of Australia’s Birds 2015” report notes that Australian Boobooks have declined in all but one region of Australia between 1999 and 2015. The report specifically stated that “This is cause for concern and further investigation is needed to understand the factors that are driving this consistent decline across regions” (BirdLife Australia, 2015). This suggests that, although somewhat resilient to the threatening processes associated with urban and agricultural development, boobooks are susceptible to some degree.

The combination of susceptibility to impacts of development and an apparent ability to persist in a variety of highly altered habitats makes the boobook an excellent model to examine the spatial distribution of the mechanisms that may be driving decline in more vulnerable predatory birds. From a purely practical perspective, their high detectability and widespread occurrence facilitates acquiring an adequate number of samples to make quantitative assessments of the prevalence of hypothesized threatening processes across areas of predominantly urban and agricultural matrix. Examination of the impacts and prevalence of the threatening processes across the three habitat types in conjunction with differences in abundance of boobooks among the three sites is necessary to understand the relative risks of these processes not only to boobooks, but to other predatory birds that are less common and more sensitive to the impacts of human development.

Maintenance of biodiversity in areas that are fragmented and heavily impacted by humans will become an increasingly important part of conservation biology as more land is developed to meet the needs of a growing human population. Understanding the mechanisms by which animal populations are impacted by fragmentation will be key to developing effective conservation strategies to allow some maintenance of biodiversity (Ricketts, 2001). To better understand these mechanisms in boobooks as a model for other predatory birds, I investigated four distinct threatening processes which I suspected to be influenced by the type of matrix between patches of remaining habitat: secondary poisoning by anticoagulant rodenticides; limitation of juvenile dispersal and impact on spatial genetic structure; resource limitation, specifically breeding site availability; and infection by the parasite Toxoplasma gondii. Simultaneous examination of the multiple threatening processes across two types of anthropogenic matrix as well as continuous natural habitat has the potential to improve our knowledge of the mechanisms by which fragmentation diminishes biodiversity and will contribute to developing strategies to mitigate these processes on a landscape scale.

I chose to investigate the relationship between anticoagulant rodenticides because relatively few studies have explored spatial aspects of anticoagulant rodenticide exposure risk. Of these studies, some were contained within specific habitat types (Cypher et al., 2014; Gabriel et al., 2012). Other studies compared exposure patterns in urban and rural habitats (Mcmillin et al., 2008; Riley et al., 2007) and found a positive correlation between

7 exposure and use of urban habitat. However, these studies have been limited to mammal species in North America and none simultaneously addressed exposure risk associated with use of agricultural systems. Additionally, prior to this study, no systematic testing for exposure to ARs had been conducted in any Australian wildlife species (Lohr, 2018; Lohr and Davis, 2018). Determining risk of exposure within agricultural systems is particularly important within Australia because of regional peculiarities in the anticoagulants used and their patterns of application (Lohr and Davis, 2018). Incidental observation of these differences prompted a literature review of the use, regulation and non-target impacts of ARs in Australia to better understand the context in which detected exposure occurred.

I simultaneously investigated spatial genetic structure in boobooks across both urban and agricultural landscapes and used band recoveries and re-sightings of banded boobooks to quantify dispersal of juveniles across fragmented habitats. Habitat fragmentation has been linked to genetic spatial structuring in Mediterranean Eagle Owls (Bubo bubo) (León-Ortega et al., 2014) and greater relatedness in urban populations of European Kestrels (Falco tinnunculus) (Riegert et al., 2010). Within Australia, urban development has been associated with numerical declines in Powerful Owls (Ninox strenua) and inbreeding between close relatives in using urban habitats (Hogan and Cooke, 2010). Accordingly, I sampled boobooks from across Western Australia to determine if habitat fragmentation was related to spatial genetic structure and genetic diversity.

I also investigated nest site limitation across different types of habitat fragmentation. Areas of continuous bushland have higher densities of tree hollows than remnant bushlands of equivalent size in urban landscapes (Davis et al., 2014; Harper et al., 2005). In agricultural remnant bushlands, hollows are being lost faster than they are being created (Saunders et al., 2014, 1982). Boobooks are obligate hollow nesters and tree hollows are a critical component of their habitat and their availability defines the borders of their continental range (Olsen and Taylor, 2001; Taylor and Canberra Ornitholgists Group, 1992). A reduction in the availability in this critical resource could potentially lead to reductions in boobook abundance. Investigating the impact of hollow availability across both urban and agricultural landscapes is important because the processes driving hollow loss vary between the two landscapes and may lead to differences in the severity of hollow loss.

The last potential threatening process I examined was infection of boobooks by the parasite Toxoplasma gondii. T. gondii is a cosmopolitan apicomplexan parasite with an extremely broad host range including all birds and mammals (Dubey, 2002) but its definitive hosts are all within the family Felidae. While T. gondii is not typically lethal in owls (Mikaelian et al., 1997) and does not appear to cause acute symptoms in experimentally infected owls (Dubey et al., 1992) but has been documented to cause mortality and serious illness in a number of native Australian marsupial species (Patton et al. 1986; Canfield et al. 1990). Raptors are susceptible to infection because of their diet and are good bioindicators of environmental prevalence of T. gondii (Love et al., 2016) particularly non-migratory owls (Gazzonis et al., 2018). As a consequence, boobooks may be a useful model for exposure in more vulnerable species. Parasite prevalence is altered by habitat fragmentation across a variety of animal taxa (Froeschke et al., 2013; King et al., 2007; Trejo-Macías et al., 2007) and previous work has demonstrated a link between T. gondii prevalence in wildlife and urban development (Barros et al., 2018). Testing seroprevalence in boobooks across unfragmented, urban, and agricultural landscapes was conducted to assess relative levels of environmental T. gondii contamination.

Consequently, my thesis focuses on the following primary objectives:

1. Critically review literature on anticoagulant rodenticide exposure in native wildlife in Australia to clarify its role as a threatening process (Chapter 2). 2. Investigate the relationship between exposure to anticoagulant rodenticides and urban and agricultural fragmentation (Chapter 3). 3. Determine if urban and agricultural fragmentation influence boobook genetic structure (Chapter 4). 4. Examine whether nest box supplementation increases site occupancy at unoccupied sites and whether this effect differs between urban and agricultural landscapes (Chapter 5). 5. Explore patterns of Toxoplasma gondii seropositivity in boobooks across the urban, agricultural, and natural landscapes (Chapter 6).

Chapter 2 Anticoagulant rodenticide use, non-target impacts and regulation: A case study from Australia

Lohr, M.T., Davis, R.A., 2018. Anticoagulant rodenticide use, non-target impacts and regulation: A case study from Australia. Sci. Total Environ. 634, 1372–1384. https://doi.org/10.1016/j.scitotenv.2018.04.069

Abstract The impacts of anticoagulant rodenticides (ARs) on non-target wildlife have been well documented in Europe and North America. While these studies are informative, patterns of non-target poisoning of wildlife elsewhere in the world may differ substantially from patterns occurring in Australia and other countries outside of cool temperate regions due to differences in the types of ARs used, patterns of use, legislation governing sales, and potential pathways of secondary exposure. Most of these differences suggest that the extent and severity of AR poisoning in wildlife may be greater in Australia than elsewhere in the world. While many anecdotal accounts of rodenticide toxicity were found – especially in conjunction with government control efforts and island eradications – no published studies have directly tested rodenticide exposure in non-target Australian wildlife in a comprehensive manner. The effects of private and agricultural use of rodenticides on wildlife have not been adequately assessed. Synthesis of reviewed literature suggests that anticoagulant rodenticides may pose previously unrecognised threats to wildlife and indigenous people in Australia and other nations with diverse and abundant reptile faunas relative to countries with cooler climates and more depauperate herpetofaunas where most rodenticide ecotoxicology studies have been conducted. To address the identified knowledge gaps we suggest additional research into the role of reptiles as potential AR vectors, potential novel routes of human exposure, and comprehensive monitoring of rodenticide exposure in Australian wildlife, especially threatened and endangered omnivores and carnivores. Additionally, we recommend regulatory action to harmonise Australian management of ARs with existing and developing global norms.

Introduction Anticoagulant rodenticides (ARs) are used worldwide in the management of introduced commensal rodents and their associated threats to crops, infrastructure, and human health (Bradbury, 2008). Baiting with ARs is also the most frequently-used method of eradicating rodents from islands and fenced areas for the purpose of preserving or reintroducing native biodiversity (Hoare and Hare, 2006). These rodenticides function by indirectly blocking recycling of vitamin K, which is a critical component in normal blood clotting in vertebrates (Park et al. 1984). ARs are often divided into first and second generation anticoagulant rodenticides based on when they were first synthesized and differences in chemical structure. Second generation anticoagulant rodenticides (SGARs) generally have higher acute toxicities than first generation anticoagulant rodenticides (FGARs) (Thomas et al., 2011). SGARS are also lethal after a single feed, unlike FGARs which require rodents to feed on them for multiple consecutive days in order to achieve a lethal effect (Erickson and Urban, 2004). During this time, rodents can continue to feed and accumulate higher concentrations of ARs (Bradbury, 2008). Retention time can vary dramatically between rodenticides but is generally highest in second generation anticoagulant rodenticides. For example, in birds, the United States EPA estimates liver retention times of 35 days for the FGAR warfarin and liver retention times of 248 days and 217 days for the SGARs bromodiolone and brodifacoum, respectively (Erickson and Urban, 2004). This long duration of SGAR persistence in liver tissues allows bioaccumulation and biomagnification in predatory species (Martínez-Padilla et al., 2016). The threat of secondary toxicity is exacerbated by behavioural changes induced in species which directly consume poisoned bait. Pre-lethal effects of ARs include reduced escape response and atypical movement in wood mice (Apodemus sylvaticus) and bank voles (Clethrionomys glareolus) (Brakes and Smith, 2005) as well as altered activity cycles and a startle response that shifted from bolting to freezing when threatened in brown rats (Rattus norvegicus) (Cox and Smith, 1992). Secondary toxicity has been demonstrated in the laboratory in a wide variety of species (reviewed in Joermann 1998) and toxicity in strict carnivores which are unlikely to eat poisoned bait is well-documented in wild animals (reviewed in Laakso et al. 2010). One study even found anticoagulant rodenticide contamination in four of four mountain lions (Puma concolor) sampled, with the deaths of two of the individuals directly attributable to acute anticoagulant intoxication (Riley et al.

2007). Lethal intoxication of an apex predator suggests substantial movement of anticoagulant rodenticides through several trophic levels and is clearly a cause for concern. Consequently, secondary poisoning of wildlife has been identified as a meaningful threat at the population level in several species (Nogeire et al., 2015; Thomas et al., 2011). The vast majority of both laboratory and field studies of non-target AR poisoning have been conducted in North America, Europe and New Zealand, but few studies have investigated secondary poisoning of wildlife in Australia, where at present, this problem is not widely recognised. The need for additional research into non-target impacts of anticoagulant rodenticides in Australia was identified as early as 1991 and such research was characterised as “required urgently” (Twigg et al., 1991). With some common predatory bird species experiencing unexplained range-wide declines (BirdLife Australia, 2015) and a suite of carnivorous dasyurid marsupials that are already threatened by disease and introduced carnivores (Burbidge and McKenzie, 1989; Woinarski et al., 2015), there is an urgent imperative to understand the role of rodenticide in the decline of susceptible wildlife species in Australia. Aims The aims of this study are to review the existing evidence for the impacts of anticoagulant rodenticides on native Australian wildlife and to highlight knowledge gaps and contextualise non-target mortality in Australia relative to other parts of the world where more comprehensive literature exists. We also sought to document the ARs currently used in Australia and to clarify the differences in legislation governing rodenticide use between Australia and a selection of other developed nations. Additionally, we highlight global literature which suggests serious knowledge gaps regarding potentially dangerous impacts of anticoagulant rodenticides on non-target wildlife and indigenous people in Australia and other nations with diverse reptile faunas. Methods Literature included in this review was obtained by searching Web of Science and Scopus databases for all articles containing the keyword “Australia” in combination with the following keywords: rodenticide, anticoagulant, brodifacoum, bromadiolone, coumatetralyl, difenacoum, diphacinone, difethialone, flocoumafen, pindone, and warfarin. Only articles containing information about the use, wildlife impacts, human exposure and regulation of

12 anticoagulant rodenticides in Australia were retained. References within these papers were searched to locate additional sources of information including PhD theses and government reports. We excluded agricultural bait development trials using baits which did not contain active ingredients, modelling of baiting regimes, therapeutic use of anticoagulants, lab toxicity trials unrelated to native Australian wildlife, government fact sheets, and other studies that did not directly involve the application of anticoagulant rodenticides or their impacts in Australia. Sources were assigned to seven categories based on their primary topic (Table 2.1). In the course of the review, major knowledge gaps relating to interactions between anticoagulant rodenticides and reptiles became apparent. To address these gaps and explore potential impacts in Australia, it was necessary to search world literature relating to reptiles and AR. We followed the same search protocol using the keywords reptile, snake, and lizard in combination with the following keywords: rodenticide, anticoagulant, brodifacoum, bromadiolone, coumatetralyl, difenacoum, diphacinone, difethialone, flocoumafen, pindone, and warfarin. Only literature relating to exposure and impacts of ARs on reptiles was examined. All searches were conducted in December 2017 and January 2018. Results and Discussion

Literature Survey We located a total of 45 publications relating to the use, impacts, and regulation of anticoagulant rodenticides in Australia (Table 2.1). The most common category of literature included 14 resources comprising 30% of all available publications and related to the documentation of island eradications of rabbits or rodents undertaken for conservation management. While eleven resources related primarily to AR impacts on non-target wildlife, none directly tested rodenticide exposure in a large number of individuals and many were reports of opportunistic observations. One publication, categorised as relating to rodenticide impacts on native wildlife, included only speculative mentions of potential poisoning (Olsen, 1996). Eight resources focused on developing AR-based methods for control of rodents, rabbits and pigs, primarily in agricultural settings. Only five studies related to laboratory testing of toxicity of ARs to non-target Australian wildlife. One tested the toxicity of the FGAR pindone to five Australian bird species (Martin et al., 1994). The

13 other four studies tested toxicity of pindone (Jolly et al., 1994) and the SGAR brodifacoum in brushtail possums (Trichosurus vulpecula) (Eason et al., 1996; Littin et al., 2002) and brodifacoum in red-necked wallaby (Macropus rufogriseus) (Godfrey, 1984) for the purpose of developing control protocols for these species in New Zealand where they are introduced pests. While toxicity literature from elsewhere in the world is likely to be useful in evaluating the risk of ARs to many Australian taxa, a lack of information on the toxicity of ARs to reptiles and marsupial carnivores prevents meaningful assessment of the potential risks posed to these groups.

Table 2.1 Numbers and categories of publications relating to anticoagulant rodenticides in Australia.

Study Type Number of Publications Island Eradications 14 Non-target Wildlife Impacts 11 Agricultural/Feral Control Trials 8 Captive Study 5 Human Exposure 4 Pindone Reviews 2 Pet Exposure 1 Total 45

Anticoagulant Exposure of Non-target Wildlife in Australia We found fifteen sources which described suspected or confirmed cases of anticoagulant rodenticide poisoning in 37 Australian wildlife species (Table 2.2). Additional cases of poisoning in carnivorous birds held in rehabilitation facilities as a consequence of encountering poisoned rodents while in care have also been reported in a Tasmanian Wedge-tailed Eagle (Aquila audax fleayi), a Grey Goshawk (Accipiter novaehollandiae), and a Tasmanian Masked Owl (Tyto novaehollandiae castanops) (Mooney, 2017) but these records were not included in Table 2.2 because the poisonings occurred in captivity. Records of wild animal poisonings occurred across the Australian Canberra Territory, the territory of Norfolk Island and all Australian states except for South Australia. One FGAR (pindone) and two SGARs (brodifacoum and bromadiolone) were implicated in the poisonings. Five mammal species, 31 bird species and one reptile species were represented in the records (Table 2.2). Three species recorded as being poisoned are listed as vulnerable (Boodie (Bettongia lesueur), Tasmanian Masked Owl (Tyto novaehollandiae castanops), and Northern Giant Petrel (Macronectes halli)) and two species are listed as endangered (Norfolk Island Boobooks (Ninox novaeseelandiae undulata) and Southern Giant Petrel

(Macronectes giganteus)). Additionally, another paper raised concern over the role that ARs might play in the decline of the Eastern Quoll (Dasyurus viverrinus), a dasyurid marsupial which is listed as endangered (Fancourt, 2016). Further research has been suggested to determine risk levels in this species but no empirical data are available on incidence of secondary toxicity or exposure rates (Fancourt, 2016). Out of the fifteen reports of wildlife poisoning, twelve were definitively related to large deployments of bait by government agencies or farmers for the purposes of island eradications, agricultural rodent control, or rabbit control (Table 2.2). Only two of the sources specifically implicated small-scale private use of rodenticides in the poisoning of wildlife (Mooney, 2017; Reece et al., 1985). Such use is largely unregulated and unmonitored and occurs in a large proportion of inhabited locations (Mooney, 2017).

1 Table 2.2 Accounts of non-target AR toxicity in Australian wildlife. *Authors do not specify how poisoning was verified

Likely Exposure Deitary Species Number Rodenticide Certainty State/Territory Source Type Category Reference Reptiles King’s skink (Egernia physical island rat kingii) 8 brodifacoum symptoms Western Australia eradication primary omnivore Bettink, 2015 Birds Norfolk Island Boobook (Ninox novaeseelandiae unspecified rat undulata) N/A brodifacoum suspected Norfolk Island control program secondary carnivore Debus, 2012 Straw-necked Ibis agricultural (Threskiornis physical mouse control invertivore spinicollis) 1 bromadiolone symptoms New South Wales trial secondary /carnivore Saunders, 1983

Barking Owl (Ninox physical connivens) 1 unknown symptoms Queensland unknown secondary carnivore Thomas & Kutt, 1997 liver analysis (unknown agricultural rat Barn Owl (Tyto alba) 1 brodifacoum concentration) Queensland control secondary carnivore Thomas & Kutt, 1997 liver analysis Lesser Sooty Owl (0.007 and agricultural rat (Tyto multipunctata) 2 brodifacoum <0.005 mg/kg) Queensland control secondary carnivore Thomas & Kutt, 1997 Masked Owl (Tyto liver analysis agricultural rat novaehollandiae) 1 brodifacoum (0.17 mg/kg) Queensland control secondary carnivore Thomas & Kutt, 1997 Southern Boobook (Ninox museum novaeseelandiae) 1 unknown record Queensland unknown secondary carnivore Thomas & Kutt, 1997 Brahminy Kite island rat (Haliastur indus) 2 pindone suspected Western Australia eradication secondary carnivore Martin et al., 1994

Brown Falcon (Falco physical private rodent berigora) 1 unknown symptoms Tasmania control secondary carnivore Mooney, 2017 Brown Goshawk physical private rodent (Accipiter fasciatus) 2 unknown symptoms Tasmania control secondary carnivore Mooney, 2017 Collared Sparrowhawk (Accipiter physical private rodent cirrocephalus) 1 unknown symptoms Tasmania control secondary carnivore Mooney, 2017 Grey Goshawk (Accipiter physical private rodent novaehollandiae) 5 unknown symptoms Tasmania control secondary carnivore Mooney, 2017 Tasmanian Masked Owl (Tyto novaehollandiae physical private rodent castanops) 12 unknown symptoms Tasmania control secondary carnivore Mooney, 2017 Tasmanian Boobook (Ninox novaeseelandiae physical private rodent leucopsis) 6 unknown symptoms Tasmania control secondary carnivore Mooney, 2017

Little Eagle (Hieraaetus morphnoides) N/A pindone suspected ACT rabbit control secondary carnivore Olsen et al., 2013 Wedge-tailed Eagle (Aquila audax) N/A pindone suspected ACT rabbit control secondary carnivore Olsen et al., 2013 Whistling Kite (Haliastur sphenurus) N/A pindone suspected ACT rabbit control secondary carnivore Olsen et al., 2013 Buff-banded Rail (Gallirallus physical island rat philippensis) 5 brodifacoum symptoms Western Australia eradication primary invertivore Palmer, 2014 Silver Gull (Larus physical island rat invertivore novaehollandiae) 7 brodifacoum symptoms Western Australia eradication both /carnivore Palmer, 2014

Pacific Golden Plover island rabbit (Pluvialis fulva) 1 brodifacoum suspected Western Australia eradication both invertivore Palmer, 2014 Ruddy Turnstone physical island rat (Arenaria interpres) 28 brodifacoum symptoms Western Australia eradication secondary invertivore Palmer, 2014 Buff-banded Rail (Gallirallus island rabbit not philippensis) 2 brodifacoum suspected New South Wales eradication specified omnivore Priddel et al., 2000 Pied Currawong island rabbit not (Strepera graculina) 1 brodifacoum suspected New South Wales eradication specified omnivore Priddel et al., 2000 Little Raven (Corvus physical residential not mellori) 1 bromadiolone symptoms Victoria rodent control specified omnivore Reece et al., 1985 Purple Swamphen (Porphyrio porphyrio physical residential not melanotus) 1 bromadiolone symptoms Victoria rodent control specified omnivore Reece et al., 1985 Brown Skua (Stercorarius antarcticus physical island rabbit Tasmania Parks and lonnbergi) 512 brodifacoum symptoms Tasmania eradication secondary carnivore Wildlife Service, 2014 Kelp Gull (Larus physical island rabbit invertivore Tasmania Parks and dominicus) 988 brodifacoum symptoms Tasmania eradication primary /carnivore Wildlife Service, 2014 Northern Giant Petrel (Macronectes physical island rabbit Tasmania Parks and giganteus) 693 brodifacoum symptoms Tasmania eradication secondary carnivore Wildlife Service, 2014 Pacific Black Duck (Anas superciliosa superciliosa) and Mallard (A. platyrhynchos physical island rabbit Tasmania Parks and platyrhynchos) 157 brodifacoum symptoms Tasmania eradication primary omnivore Wildlife Service, 2014 Southern Giant Petrel (Macronectes physical island rabbit Tasmania Parks and halli) 38 brodifacoum symptoms Tasmania eradication secondary carnivore Wildlife Service, 2014 Unknown Bird 5 brodifacoum physical Tasmania island rabbit not Tasmania Parks and

symptoms eradication specified Wildlife Service, 2014 Unknown giant petrel (Macronectes physical island rabbit Tasmania Parks and sp.) 31 brodifacoum symptoms Tasmania eradication secondary carnivore Wildlife Service, 2014 Australian Ringneck (Barnardius zonarius) N/A pindone suspected Western Australia rabbit control primary herbivore Twigg et al., 1999 Brahminy Kite (Haliastur indus) N/A pindone suspected Western Australia rabbit control secondary carnivore Twigg et al., 1999 Crested Pigeon (Ocyphaps lophotes) N/A pindone known* Western Australia rabbit control primary herbivore Twigg et al., 1999 Grass Owl (Tyto agricultural rat longimembris) 1 brodifacoum liver analysis Queensland control secondary carnivore Young & De Lai, 1997 Masked Owl (Tyto physical agricultural rat novaehollandiae) 1 brodifacoum symptoms Queensland control secondary carnivore Young & De Lai, 1997 Rufous Owl (Ninox physical agricultural rat rufa) 2 brodifacoum symptoms Queensland control secondary carnivore Young & De Lai, 1997 Mammals southern brown bandicoots (Isoodon obesulus) N/A pindone liver analysis Western Australia rabbit control primary omnivore Twigg et al., 1999 swamp wallaby (Wallabia bicolor) N/A pindone known* New South Wales rabbit control primary herbivore Twigg et al., 1999 western grey kangaroo (Macropus fuliginosus) N/A pindone known* Western Australia rabbit control primary herbivore Twigg et al., 1999 brushtail possum (Trichosurus physical not vulpecula) 7 unknown symptoms Queensland unknown specified omnivore Grillo et al., 2016 boodie (Bettongia population island rat lesueur) 20-50 pindone eradicated Western Australia eradication primary herbivore Morris, 2002

2 In addition to accounts of wildlife poisoning, we also located published accounts 3 suggesting population-level effects of rodenticide toxicity on carnivorous birds in Australia. 4 Olsen (1996) listed the use of rodenticides in areas of palm cultivation as a potential 5 contributing factor in the decline of Norfolk Island Boobooks (Ninox novaeseelandiae 6 undulata x novaeseelandiae). Young and De Lai (1997) observed a correlation between 7 declines in owl abundance and the use of “Klerat®”a brodifacoum-based rodenticide in 8 sugar cane fields in north Queensland and documented one confirmed and several 9 suspected cases of brodifacoum poisoning in owls (James, 1997). A subsequent report 10 noted three additional cases of owls in Queensland testing positive for brodifacoum 11 residues (0.007 mg/kg, <0.005mg/kg, and 0.17mg/kg ) in the 1990s and two museum 12 specimens of Southern Boobooks (Ninox novaeseelandiae) with rodenticide poisoning listed 13 as their cause of death in the collection notes (Thomas and Kutt, 1997). One of the two 14 specimens, while alive showed symptoms of AR poisoning including “bleeding from the 15 nasal passages; loss of muscle co-ordination; lethargy including drooping head and eyes; 16 and generally poor and dirty condition” (Thomas and Kutt, 1997). The report reviewed 17 several other factors which could potentially have impacted owl populations in the area and 18 came to the conclusion that there was “significant potential for secondary poisoning of owls 19 to occur in Queensland sugarcane as a result of the use of Klerat®” (Thomas and Kutt, 1997). 20 Crop Care Australia later deregistered Klerat® for use in sugar cane fields over concerns 21 relating to secondary poisoning (Twigg et al., 1999).

22 An unpublished PhD dissertation examined dynamics of secondary poisoning of 23 avian predators associated with sugar cane fields in Queensland and concluded that the 24 coumatetralyl-based product used to control rats did not pose a threat to predatory birds 25 (Ward, 2008). This conclusion was based largely on the low relative use of canefields for 26 foraging by predatory birds, the low concentration of coumatetralyl in rats captured outside 27 of canefields, and the low toxicity and persistence of coumatetralyl relative to second 28 generation anticoagulant rodenticides (Ward, 2008). Unfortunately, no predatory birds in 29 the treated areas were directly tested for rodenticide exposure. A lack of detection of 30 coumatetralyl in Southern Boobooks in Western Australia as part of an ongoing study 31 supports the low probability of secondary toxicity in raptors for this rodenticide.

32 Pindone has been implicated as a factor driving the decline of Little Eagle (Hieraaetus 33 morphnoides) numbers in and around Canberra (Olsen et al. 2013). Breeding pairs of Little 34 Eagles disappeared from areas baited with pindone while pairs in areas baited with 1080 or 35 not baited at all persisted (Olsen et al. 2013). The high susceptibility of Wedge-tailed Eagles 36 to pindone in laboratory tests (Martin et al. 1994) lends credibility to the hypothesis that 37 pindone could be responsible. Unfortunately, no direct testing of Little Eagles suspected of 38 being poisoned was conducted to confirm pindone exposure and rule out other ARs from 39 residential and commercial sources.

40 Recently, a study in Tasmania examined probable rodenticide poisoning in predatory 41 birds. Six species (Table 2.2) showed signs of anticoagulant rodenticide poisoning when 42 dissected (Mooney, 2017) but the rodenticides responsible were not determined or 43 quantified. As part of this study, thirteen predatory bird species were ranked by risk of 44 rodenticide exposure according to four natural history parameters: relative metabolic 45 speed, dietary habits influencing consumption of contaminated tissues, relative preference 46 for rodents, and willingness to forage near anthropogenic structures (Mooney, 2017). 47 Development of a more statistically robust predictive model using similar natural history 48 parameters to examine risk of rodenticide exposure in a wider range of predatory species 49 would be an extremely useful step toward assessing likely population level impacts on 50 wildlife in Australia. Incorporating variables relating to seasonal dietary shifts and home 51 range size could potentially improve future models.

52 The overall lack of attention within Australia to what is perceived as a potentially 53 serious threatening process for native carnivores in many other parts of the world suggests 54 the need for Australian studies which examine potential impacts on native fauna in a 55 quantitative and comprehensive manner. Susceptibility of marsupial carnivores is 56 particularly poorly understood and should be a focus of future research. Furthermore, a 57 surveillance program should be in place in areas of high AR use, to monitor any dead wildlife 58 for a cause of death. Most of the studies we used in this review did not sample animals and 59 thus were not able to confirm suspicions of death due to rodenticide poisoning.

60 4.3 Governance and legislation of rodenticide use

61 At present, no information is available on the volume of sales or application of ARs in 62 Australia. Reporting for all poisons intended to control vertebrates indicates that 222 63 different products are currently registered with a total sales reaching $18,601,875.00 in the 64 2015-2016 fiscal year (Australian Pesticides and Veterinary Medicines Authority, 2017a). 65 Nine anticoagulants are currently approved for vertebrate pest control in Australia (McLeod 66 and Saunders 2013). At present, all nine are listed as Schedule 6 substances (see Appendix 67 2.A for schedule meanings) in Australia (Australian Government Department of Health: 68 Therapeutic Goods Administration, 2017) (Table 2.3) and are legally allowed to be sold 69 directly to the public and do not require government permits for purchase or use. In some 70 cases, more concentrated formulations of SGARs are listed as Schedule 7 substances and are 71 restricted to licensed pesticide applicators (Australian Government Department of Health: 72 Therapeutic Goods Administration, 2017) while products containing low concentrations of 73 some FGARs are registered as schedule 5 substances which require only simple warnings 74 and safety directions for public sale (Table 2.3). The FGAR diphacinone is currently approved 75 as an active ingredient but has no products registered with the APVMA after July 2016 76 (Australian Pesticides and Veterinary Medicines Authority, 2017b). However, remaining 77 stock can still be used for 12 months following a stopped registration (Commonwealth of 78 Australia, 1994) and MSDS sheets obtained from a pest management contractor seem to 79 indicate that at least one diphacinone product is still in use at present. The APVMA has 80 prioritised a review of the status of all SGARs currently approved in Australia (brodifacoum, 81 bromadiolone, difenacoum, difethialone, and flocoumafen) citing concerns over public 82 health, worker safety, and environmental safety (Australian Pesticides and Veterinary 83 Medicines Authority, 2015)

84 Table 2.3 Anticoagulants currently approved for vertebrate pest control in Australia. Some anticoagulants are assigned different schedules dependant on formulation. *Some disagreement 85 exists as to whether these should be treated as first or second generation anticoagulants †Warfarin is used therapeutically in humans as a blood thinner.

Acute Oral Schedule (See LD (Rattus Approved Anticoagulant Chemical Class Generation 50 LD Reference Appendix 2.A) norvegicus) 50 Target Species mg/kg 6 (0.25 per cent or brodifacoum hydroxycoumarins second Godfrey 1985 mice and rats less) or 7 0.27 6 (0.25 per cent or bromadiolone hydroxycoumarins second 0.57-0.75 Meehan 1978 less)or 7 mice and rats 5 ( 0.05 per cent or coumatetralyl hydroxycoumarins first less), 6 (1 per cent Dubock and Kaukeinen 1978 or less), or 7 16.5 mice and rats 6 (0.25 per cent or difenacoum hydroxycoumarins second Bull 1976 less) or 7 1.8-3.5 mice and rats hydroxyl-4- 6 (0.0025 per cent difethialone second Lechevin and Poche 1988 benzothiopyranones or less) or 7 0.27-0.69 mice and rats approval diphacinone indandiones first* 6 1.93-2.7 Fisher et al. 2003 expired 6 (0.005 per cent or flocoumafen hydroxycoumarins second 0.25-0.56 Lund 1988 less) or 7 mice and rats pindone indandiones first* 6 75-100 Fisher et al. 2003 rabbits 4†, 5 (0.1 per cent warfarin hydroxycoumarins first Fisher et al. 2003 or less), or 6 3.3 mice and rats

86 Increasing concerns over risks to the health and safety of humans and pets and 87 impacts on non-target wildlife have prompted stricter regulation of anticoagulant 88 rodenticides – particularly SGARs – in several developed nations. While rodenticide 89 legislation is often complex and varies substantially between countries, the trend is toward 90 stricter legislation than currently exists in Australia. In the United States, SGARs are 91 restricted to licensed pesticide applicators, only allowed to be used indoors, and are 92 required to be placed in containers which exclude children and pets (Bradbury, 2008). 93 Similar requirements were subsequently implemented in Canada (Health Canada: Pest 94 Management Regulatory Agency, 2010). A somewhat different approach is taken in the UK, 95 where SGARS are licensed for outdoor use but an industry taskforce has been established to 96 monitor both rodenticide applicator usage patterns and breeding success and SGAR residues 97 in the livers of one sentinel species – Barn Owls (Tyto alba) – to determine the impacts of 98 this legislative change on exposure rates (Shore et al., 2016). These alternative models of 99 AR regulation and the direction they represent in evolving global norms should be 100 considered when evaluating current Australian regulations.

101 Given the changes in legislation governing the use of ARs in other developed nations 102 and demonstrated impacts on human health and wildlife populations overseas, we support 103 the ongoing review of the use and scheduling of SGARs in Australia by the APVMA. In 104 Australia, AR poisoning has been documented in pets (Robertson et al., 1992) and humans 105 (Osborne et al., 2017), particularly children (Ozanne-Smith et al., 2001; Parsons et al., 1996; 106 Reith et al., 2001). Roughly 1,400 human exposures to ARs per year are recorded by Poison 107 Information Centres in Australia (Australian Pesticides and Veterinary Medicines Authority, 108 2015). Removal of SGARs from retail sale to the public by listing all SGARs as schedule 7 109 poisons and implementing stricter requirements that baits be used only indoors and placed 110 in a manner that makes them inaccessible to children and pets will help to bring Australian 111 practices closer to emerging global norms and best practices. These actions are likely to 112 help to mitigate human health and safety risks and exposure in non-target wildlife. Critical 113 evaluation of whether these practices are effective will require long-term monitoring of AR 114 residues in appropriate sentinel species – as practiced in the UK – before and after any 115 regulatory changes are implemented. Ongoing research into exposure patterns in Southern

116 Boobooks will provide valuable baseline data for a widely-distributed sentinel species if the 117 suggested regulatory changes are implemented.

118 Current Uses in Australia

119 Agricultural 120 In Australian agriculture, ARs are primarily used in asset protection around 121 infrastructure and grain storage areas and many first and second generation products are 122 licensed for these purposes. In the past, several trials have been conducted on broadscale 123 application of rodenticides in Australian cropping systems.

124 Brown & Singleton (1998) found aerial distribution of brodifacoum-based baits 125 effective at controlling mice in wheat fields in South Australia in a field trial and the authors 126 suggested that application according to guidelines was unlikely to cause substantial non- 127 target mortality. However, mice were observed to be active during the day following the 128 baiting, which the authors acknowledged could increase the risk of secondary poisoning in 129 predatory species (Brown and Singleton, 1998). To our knowledge, aerial distribution of 130 brodifacoum baits in wheat crops has never been implemented on an operational basis in 131 Australian agriculture.

132 Several trials of bromadiolone efficacy in controlling mouse plagues have been 133 conducted in agricultural crops in Australia. In the earliest of these studies, aerial 134 application was used to distribute bromadiolone bait directly into sunflower crops in New 135 South Wales (Saunders, 1983). Bromadiolone was identified as the most promising of the 136 three toxicants tested but the authors noted concern over bromadiolone’s slow method of 137 action potentially facilitating secondary poisoning of predators selecting for poisoned mice 138 (Saunders, 1983). One Straw-necked Ibis (Threskiornis spinicollis) was found dead of 139 apparent rodenticide poisoning after having consumed 6-10 mice in an area where 140 bromadiolone had been aerially applied as part of a trial to control mice in sunflower crops 141 (Saunders, 1983). In a subsequent study, wheat laced with bromodialone was applied a 142 single time in bait stations in soybean crops in New South Wales (Twigg et al., 1991). The 143 study did not search for or detect any mortalities in non-target wildlife but cautioned that 144 “The risks to non-target species and of contaminating primary produce posed by broad-scale 145 use of rodenticides would need to be assessed fully before these chemicals could become

146 an integral part of farm management. In Australia, such data are sparse and research is 147 required urgently” (Twigg et al., 1991). The only subsequent available study on broad-scale 148 use of bromadiolone in agriculture used a fertiliser spreader to apply four treatments of 149 wheat laced with bromadiolone to “refuge habitat, channel banks, fence lines, non-arable 150 land and road verges” within 200m of soybean crops in New South Wales but failed to 151 demonstrate significant reductions in crop damage (Kay et al., 1994). It does not appear 152 that non-target exposure was evaluated as part of this study.

153 Contrary to the warning issued by Twigg et al. (1991), which cautioned a more 154 complete assessment of non-target impact prior to the broad-scale use of ARs in agriculture, 155 under some circumstances, ARs are or have been used in or adjacent to crops to control 156 mice and rats. During mouse plagues, temporary registrations for the use of bromadiolone 157 have been issued for use in wheat crops in Victoria in 1984, perimeter baiting of oilseed 158 crops in New South Wales in 1984-1985, and in soybean crops in New South Wales in 1989 159 (Twigg et al., 1991). Expired permits issued to allow the baiting of crop perimeters with 160 bromadiolone show valid periods between 16 September 1999 and 31 December 1999 161 (PER3031); 06 December 2006 and 30 March 2009 (PER9543); and 31 March 2009 and 30 162 June 2016 (PER11331) (Australian Pesticides and Veterinary Medicines Authority, 2017b). 163 There are no current permits for the use of bromadiolone in perimeter baiting around crops 164 but a current New South Wales government factsheet and web page state that 165 bromadiolone bait can be prepared by the Livestock Health and Pest Authority (LHPA) for 166 availability to farmers in perimeter baiting around crops (New South Wales Department of 167 Primary Industries, 2011; New South Wales Government: Department of Primary Industries, 168 2017).

169 The SGAR brodifacoum was also previously applied broadscale in sugar cane fields in 170 Queensland (Young and De Lai, 1997) but the registration for that use has since been 171 revoked over concerns about mortality in non-target wildlife (Twigg et al., 1999). The use of 172 brodifacoum in this context has largely been replaced by the use of the FGAR coumatetralyl. 173 Research on non-target impacts of coumatetralyl in sugar cane fields demonstrated low risk 174 of secondary toxicity (Ward, 2008). Coumatetralyl is currently registered for use in 175 pineapple, macadamia, and sugar cane crops in all states and territories (Australian 176 Pesticides and Veterinary Medicines Authority, 2017b).

177 Published literature and official accounts may seriously underestimate the usage of 178 ARs in cropping systems in Australia. A study of second generation anticoagulant use in 179 agricultural systems in Northern Ireland found that total compliance with best practice 180 application methods was rare and lack of compliance probably facilitated greater risk of 181 secondary toxicity to native wildlife (Tosh et al. 2011). Within Australia, landowners have 182 requested that a government agency provide them with pindone with the intention of using 183 pindone to reduce kangaroo abundance in contravention of its label (Twigg et al., 1999). 184 Many ARs are readily available in hardware and agricultural supply stores in Australia 185 without a permit and the potential for use contrary to labelling restrictions is high. A better 186 understanding of current legal and illegal usage of ARs in agriculture is necessary to 187 determine the likelihood of secondary poisoning of non-target species in agricultural 188 systems.

189 Conservation 190 ARs have a long history of use on islands and in fenced reserves worldwide for 191 eradication of rodents for conservation purposes. At present, application of ARs is the only 192 effective way of removing introduced rodents from islands larger than 5ha for conservation 193 purposes (Campbell et al., 2015). Many successful and well-documented eradications of 194 introduced rodents and rabbits have been conducted in Australia using ARs (Bettink, 2015; 195 Burbidge, 2004; Cory et al., 2011; Dunlop et al., 2015; Meek et al., 2011; Morris, 2002; 196 Priddel et al., 2000; Tasmania Parks and Wildlife Sevice, 2014). Pindone was used in some 197 early eradications but its use has largely been supplanted by brodifacoum (Burbidge and 198 Morris, 2002) and bromadiolone (Meek et al., 2011). Reviews of island eradications have 199 been conducted for New South Wales (Priddel et al., 2011) and Western Australia (Burbidge 200 and Morris, 2002).

201 During the course of some eradications, high levels of non-target mortality and 202 poisoning of species listed under the Australian Environment Protection and Biodiversity 203 Conservation Act 1999 have been documented. In one instance, boodies (Bettongia lesueur) 204 (listed as vulnerable) were accidentally eradicated on Boodie Island along with the intended 205 target, black rats (Rattus rattus) (Morris, 2002). An eradication of black rats was proposed 206 for Woody Island in Western Australia but was halted when the rats on the island were 207 subsequently identified as a native species (Rattus fuscipes) (Burbidge et al., 2012). During

208 the successful eradication of rabbits, black rats, and mice (Mus musculus) on Macquarie 209 Island, concerns were expressed by the public and government authorities over the 210 observed mortality of 2,424 individuals from several seabird and waterfowl species, 211 presumably related to the use of the SGAR brodifacoum (Tasmania Parks and Wildlife 212 Sevice, 2014). While some species, especially Northern Giant Petrels (listed as vulnerable) 213 experienced substantial population-level declines as a result of the baiting, the reductions 214 were expected to be temporary and removal of introduced mammals has already facilitated 215 improved population parameters in a number of seabird species (Tasmania Parks and 216 Wildlife Sevice, 2014). Endangered Southern Giant Petrels were also lethally poisoned 217 during the course of this eradication (Tasmania Parks and Wildlife Sevice, 2014). Collateral 218 damage to non-target species may be acceptable and necessary in some situations but more 219 careful consideration and planning are required to avoid poor outcomes which have 220 occurred or been narrowly averted during rodent eradications in the past. In some 221 instances, bait boxes modified to exclude native fauna may decrease the incidence of 222 primary of non-target wildlife AR exposure during eradication attempts (Moro, 2001). Use 223 of biological control agents prior to baiting can also increase the probability of success and 224 reduce the volume of poison needed to remove target animals (Priddel et al., 2000). Close 225 monitoring of non-target mortality during and after island eradications is necessary to 226 properly assess the relative benefit to native biodiversity.

227 In Australia, ARs have also been tested as a method to control feral pigs for 228 conservation purposes and reduction of agricultural threats. Trials using the FGAR warfarin 229 were conducted in New South Wales (Choquenot et al., 1990; Saunders et al., 1990) and the 230 Australian Capital Territory (McIlroy et al., 1989). While two of the three trials found the 231 use of warfarin to be highly effective, this method does not appear to have been put into 232 practice due to concerns over animal ethics, non-target exposure, and a shift toward the use 233 of 1080 baits for pig control (Cowled et al., 2008). However, the use of warfarin to control 234 feral pigs in Australia has been recommended in the published literature as recently as 2014 235 (McIlroy, 2014). While warfarin is unlikely to cause secondary poisoning in exposed wildlife, 236 the risk of primary poisoning to wildlife consuming bait intended for pigs is likely too high to 237 warrant the use of this method of control.

238 Residential and Commercial 239 Patterns of residential and commercial use of ARs in Australia are poorly known. At 240 present, the Australian Pesticide and Veterinary Medicine Association (APVMA) lists seven 241 ARs (two FGARs and five SGARs) as registered for use in Australia in commercial and 242 residential settings (Table 2.3). We have observed two FGARs (warfarin and coumatetralyl) 243 and three SGARs (brodifacoum, bromadiolone, and difenacoum) available for purchase by 244 the public at retail outlets in Western Australia. The SGARs flocoumafen and difethialone 245 are also used by commercial pest control companies in residential and commercial settings. 246 Residues of both have been detected in native wildlife in Western Australia. Patterns of 247 availability to unlicensed individuals are similar to those in the UK where three FGARs and 248 five SGARs are registered for use and are not restricted to licensed applicators (Shore et al., 249 2016). However, regulations governing AR use are substantially more restrictive in some 250 other industrialized countries. In the US, three FGARs are permitted for use by the public 251 but all four registered SGARs are restricted to use by licensed pesticide applicators 252 (Bradbury, 2008). Similarly, in Canada the public has access to three FGARs and licensed 253 contractors may use an additional three SGARs (Health Canada: Pest Management 254 Regulatory Agency, 2010).

255 The lack of available data on the quantities of ARs used in domestic and commercial 256 settings and the locations where they are used makes it nearly impossible to gauge the 257 potential non-target impacts of these products. Only two publications directly implicate 258 private use of rodenticides in non-target mortality in Australia. In the most definitive 259 example, brodifacoum was implicated in the deaths of a Purple Swamphen (Porphyrio 260 porphyrio melanotus) and Little Raven (Corvus mellori) which showed signs of AR poisoning 261 after baiting in a residential area (Reece et al., 1985). In Tasmania, residential and small- 262 scale agricultural baiting is thought to have been the source of ARs responsible for the 263 suspected lethal poisonings of 27 individuals from six raptor species (Mooney, 2017). Given 264 that use of rodenticides in conservation and agricultural contexts is relatively limited and 265 only occurs periodically, the total amount deployed in residential and commercial settings is 266 likely to be far greater. Accordingly, overseas studies on rodenticide exposure in bobcats 267 (Lynx rufus) in America (Riley et al., 2007) and a variety of bird and mammal species in Spain 268 (López-Perea et al., 2015) indicate a spatial correlation between population density and AR

269 exposure in wildlife. Collection of basic information on the quantities of ARs sold to private 270 residents and pest control contractors by locality coupled with systematic testing of wildlife 271 populations across different land-use types will be essential in assessing the risks posed to 272 non-target wildlife by residential and commercial use of ARs.

273 Unique Considerations in Australia

274 Pindone 275 Unlike other ARs used in Australia, the SGAR pindone has received more scrutiny and 276 has been the focus of a greater body of research because of its longer history of use and 277 large scale of use in rabbit control. At present, it is only registered for use in Australia and 278 New Zealand (P. Fisher, Brown, & Arrow, 2015; Twigg et al., 1999) and, as a consequence, 279 has received little attention by researchers elsewhere in the world. Efficacy trials for rabbit 280 control were conducted in Western Australia in 1971-1975 (Oliver et al., 1982) and 1981- 281 1982 (Robinson and Wheeler, 1983). Pindone was registered in Western Australia for rabbit 282 control in 1984 and was subsequently registered for the same use in all other Australian 283 states (Twigg et al., 1999). Pindone was registered for use in New Zealand in 1992 (Twigg et 284 al., 1999). In Australia, pindone is used in rabbit control primarily in areas where the use of 285 sodium fluoroacetate (1080) is deemed to pose too great a risk to humans and pets 286 (Department of Agriculture and Food Western Australia, 2015). Such areas include “market 287 gardens, golf courses, hobby farms, around farm buildings” (Twigg et al., 1999) and 288 bushlands adjacent to populated areas. In the past, it has also been used in island 289 eradications of rabbits and rodents prior to being largely replaced by brodifacoum (Burbidge 290 and Morris, 2002; Priddel et al., 2011).

291 Pindone use in Australia has been the subject of extensive review (National 292 Registration Authority For Agricultural and Veterinary Chemicals, 2002; Twigg et al., 1999) 293 prompted by public concern over reports of lethal poisoning of non-target species (Table 294 2.2). As a consequence, additional restrictions were placed on the sale of pindone 295 concentrates and labelling was required to include a “statement not to lay baits in the 296 vicinity of native animal habitat” (National Registration Authority For Agricultural and 297 Veterinary Chemicals, 2002).

298 At present, little is known about the effects of pindone on non-target species. 299 Pindone has been shown in laboratory tests to have varying effects on different native 300 Australian bird taxa (Martin et al. 1994). Wedge-tailed Eagles were more susceptible than 301 other species tested but Common Bronzewings (Phaps chalcoptera) and other granivores 302 were also noted to be at high risk of poisoning due to direct consumption of poisoned grain 303 (Martin et al. 1994). Despite the authors’ recommendation for field studies of impacts on 304 Wedge-tailed Eagles and other raptors (Martin et al., 1994), to the best of our knowledge, 305 no further study on this topic has been conducted in Australia.

306 The repeated use of an anticoagulant in natural areas to control but not eradicate 307 rabbits appears to be unique to Australia and New Zealand. The repeated pattern of use in 308 the same areas may pose a serious long-term threat to susceptible wildlife populations. This 309 may be especially problematic for long-lived species with low reproductive rates which are 310 unable to sustain low levels of additive mortality. The potential link between pindone 311 baiting and the decline of Little Eagles in Canberra (Olsen et al., 2013) exemplifies this 312 concern. However, an ongoing study of rodenticide exposure in Southern Boobooks has not 313 detected any pindone residue in samples tested to date despite testing of samples obtained 314 in areas where pindone baiting has occurred. Differences in diet, territory size, and 315 metabolism could account for this lack of detection. In some instances, reduction of prey 316 abundance via ARs could potentially drive declines in predatory species rather than direct 317 ARs toxicity. However, in the instance of Little Eagles in Canberra, this does not appear to 318 be the case, as the decline of Little Eagle abundance was independent of rabbit abundance 319 (Olsen et al., 2013). Additional research into the sensitivity of Australian fauna to pindone 320 and the population impacts of different patterns of use are necessary to determine the 321 extent and severity of impacts on non-target fauna. At minimum, the continued use of 322 pindone to control rabbits in bushland areas needs to be evaluated as to whether it 323 provides a net benefit or detriment to the conservation of native biodiversity.

324 Human consumption of rabbits is common in agricultural areas and may facilitate 325 some risk of human exposure to pindone. Risk of substantial human exposure is reduced by 326 the fact that livers are not typically consumed. However, pindone has been demonstrated 327 to accumulate in fat tissue in rabbits at similar concentrations to liver tissue (Fisher et al., 328 2015). Some discussions of risk of human exposure to ARs via ingestion of contaminated

329 game meats have suggested that cooking prior to consumption might reduce AR exposure 330 through degradation of the relevant chemicals (Eisemann and Swift, 2006). Conversely, 331 subsequent empirical research demonstrated that, at least in pig tissues contaminated with 332 diphacinone, cooking did not substantially reduce AR concentration (Pitt et al., 2011). While 333 we consider the risk of pindone poisoning associated with human consumption of wild 334 rabbits to be low due to its relatively short half-life and low acute toxicity, as a minimum 335 precaution we recommend adhering to established 5 week withholding period for livestock 336 exposed to pindone (Twigg et al., 1999).

337 Reptiles 338 We found only one example of documented or suspected lethal AR poisoning of 339 reptiles in Australia (Bettink, 2015) in the course of our literature search. A further 340 investigation of international literature revealed serious gaps in knowledge relating to 341 impacts of ARs on reptiles and their potential role as vectors to higher trophic levels. In 342 combination, the few existing published accounts suggest that some reptiles may be more 343 resistant to anticoagulant rodenticides than birds or mammals. As a consequence, 344 developing a better understanding of how reptiles are impacted by AR exposure and their 345 potential as vectors to more vulnerable taxa will be critical to evaluating the ecotoxicology 346 of ARs in areas of the world where reptiles are a substantial component of biodiversity.

347 The mechanisms by which carnivorous birds and mammals are exposed to ARs have 348 not been widely researched (Elliott et al., 2014). The few studies investigating AR exposure 349 in intermediate vectors tend to focus on insects (Masuda et al., 2014), and small mammals 350 (Brakes and Smith, 2005) as potential vectors (Elliott et al., 2014) with the vast majority of 351 work focusing on target and non-target small mammals (Hoare and Hare, 2006). Because 352 most of these studies have been conducted in temperate areas of Europe or North America, 353 they may not be representative of dominant exposure pathways in tropical and warm arid 354 areas of the world. In areas where reptiles are more diverse and abundant, reptiles may act 355 as an important pathway for transmission of ARs through terrestrial food webs because of 356 their increased relative importance as prey items for carnivores at higher trophic levels 357 (Hoare and Hare, 2006). Furthermore, in ecosystems with a high predominance of 358 carnivorous reptiles e.g. snakes, monitor lizards and large skinks, there may be a direct bio-

359 accumulation effect when reptiles prey on rats or mice directly, or on other reptiles, leading 360 to a negative impact on larger-bodied reptiles (Bishop et al., 2016; Olsson et al., 2005).

361 Reptiles make up a substantial proportion of the prey base of some carnivores in 362 Australia (Doherty et al., 2015; Paltridge, 2002) and comprise >80% of the biomass in the 363 diets of some predatory bird species (Aumann, 2001). Reptile diversity and abundance is 364 substantially higher in Australia than in Europe and North America (Roll et al., 2017) where 365 secondary anticoagulant rodenticide exposure has been more comprehensively assessed in 366 native fauna. As a consequence, understanding patterns of exposure in reptiles and their 367 capacity to transmit ARs to higher trophic levels is critical to understanding ecosystem level 368 AR exposure in Australia and other countries with high reptile abundance. Only a few 369 studies have investigated the mechanisms and ramifications of AR exposure in reptiles 370 (Hoare and Hare, 2006). In one instance, the SGAR brodifacoum was detected in Pinzón lava 371 lizards (Microlophus duncanensis) up to 850 days after baiting of an uninhabited island with 372 no other rodenticide sources (Rueda et al., 2016). Long duration of AR persistence in lava 373 lizards could be a consequence of recursive exposure from consumption of invertebrates 374 feeding on reptile faeces containing AR residue, low elimination rates by lizards, or slow 375 decomposition leading to prolonged availability of bait (Rueda et al., 2016). Subsequent 376 deaths of 22 Galapagos hawks (Buteo galapagoensis) showing signs of rodenticide toxicity 377 were attributed to secondary poisoning resulting from consumption of lava lizards, as was 378 the death of a short-eared owl (Asio flammeus) found dead with lethal concentrations of 379 brodifacoum present in its liver 773 days after baiting (Rueda et al., 2016). If other reptile 380 species are also capable of vectoring lethal levels of rodenticide to higher trophic levels for 381 greater than two years after initial exposure, the threat of secondary poisoning to 382 carnivorous birds and mammals in regions of the world with diverse and abundant 383 herpetofaunas may be severely underestimated.

384 High tolerance to AR exposure may also increase the efficacy of reptiles as vectors of 385 ARs to higher trophic levels. At least some reptiles appear to be substantially more resistant 386 to AR toxicity than birds or mammals (Weir et al., 2015). An acute oral LD50 of 550 μg/g 387 was determined for the AR pindone in Western fence lizards (Sceloporus occidentalis) (Weir 388 et al., 2015). No LD50 was determined for the SGAR brodifacoum because all western fence 389 lizards tested survived the highest does of 1,750 μg/g (Weir et al., 2015). Both LD50s are

390 three to five orders of magnitude higher than in most bird and mammals species tested 391 (Laakso et al., 2010). Similarly, when prairie rattlesnakes (Crotalus viridis) were fed three 392 laboratory mice poisoned with bromadiolone over the course of three weeks, none of the 393 snakes died or showed signs of rodenticide toxicity in the 30 days following the treatment 394 despite consuming more mg/Kg brodifacoum than the LD50s established for several 395 mammal species in the same study (Poché, 1988). Pitt et al. (2015) examined brodifacoum 396 residues in 112 geckoes (Lepidodactylus lugubris and Hemidactylus frenatus) collected on 397 Palmyra Atoll after rat control operations. They noted a peak concentration of 0.067 μg/g 398 and detectable concentrations at about half of this rate were still noted 60 days post-baiting 399 Pitt et al., 2015). Pitt et al. (2015) concluded that geckos were unlikely to experience 400 mortality but on islands where secondary predators existed, there could be some 401 ecosystem-wide impacts. Similarly, bungarras or Gould’s goannas (Varanus gouldii) were 402 observed consuming rats poisoned with brodifacoum during an eradication in the 403 Montebello Islands of Western Australia, but did not appear to experience adverse effects 404 (Burbidge, 2004). If a tolerance for rodenticides exists across multiple reptile taxa, reptiles 405 may be more effective at concentrating and transmitting ARs to higher trophic levels than 406 the small mammals which have been more commonly examined as potential vectors of ARs 407 to higher trophic levels.

408 Conversely, in some instances, apparent susceptibility of some reptile species to ARs 409 has been observed or hypothesized. In Australia, the single documented account of lethal 410 AR toxicity in reptiles involved the direct ingestion of brodifacoum baits by King’s skinks 411 (Egernia kingii) during a rat eradication on Penguin Island in Western Australia (Bettink, 412 2015). Eight of the skinks were found dead and exhibited haemorrhage associated with AR 413 toxicity and several others were treated with vitamin K and released (Bettink, 2015). 414 Subsequent analysis revealed a concentration of 1.3 mg/kg in the liver of one of the dead 415 skinks (Bettink, 2015). This liver concentration is well above minimum lethal thresholds 416 suggested for many bird and mammal species so it is difficult to infer relative susceptibility 417 of King’s skinks from this event. Sánchez-Barbudo et al. (2012) documented the death of a 418 horseshoe whip snake (Hemmorrhois hippocrepis) due to flocoumafen used to protect a 419 seabird colony. A number of anecdotal accounts of lethal AR poisoning have also been 420 reported in skinks and geckos (Wedding et al., 2010). Susceptibility of goannas in Australia

421 to poisoning with brodifacoum has also been suggested (James, 1997), although it appears 422 that this only considers the likelihood of exposure due to carrion being a component of their 423 diet rather than an actual vulnerability to the effects of brodifacoum. The lack of observed 424 mortality in some reptile species may be due to a delayed onset of effects relative to birds 425 and mammals. This possibility is supported by the observation of the deaths of six 426 Galápagos land iguanas (Conolophus subcristatus) more than two months after their island 427 was baited with brodifacoum to control rats. Merton (1987) described a similar incident in 428 which Telfair’s skinks (Leiolopisma telfairii) were found dead three to six weeks after AR bait 429 was used on Round Island, Mauritius. The delay in mortality was presumed to be a result of 430 some physiological difference between reptiles and bird and mammals (Merton, 1987). If 431 some reptiles are susceptible to AR poisoning but exhibit substantially delayed mortality, 432 they may be extremely effective vectors to vulnerable species in higher trophic levels if they 433 are able to ingest higher levels of rodenticide over the pre-lethal period and if mortality is 434 preceded by behaviours which increase the likelihood of predation. Laboratory toxicity tests 435 are needed across a representative suite of reptile taxa to resolve questions around the 436 dangers posed to reptiles by ARs and the capacity of reptiles to vector ARs to higher trophic 437 levels. Extensive testing of wild reptiles would be useful in assessing exposure rates and 438 ecological impacts of reptile exposure to ARs.

439 Primary consumption of ARs by reptiles through direct consumption of baits 440 intended for rodents also requires additional evaluation as a source of AR contamination in 441 terrestrial ecosystems. In captive trials, some but not all skinks (Oligosoma maccanni) 442 consumed or licked pindone bait, with increased consumption when the bait was wet 443 (Freeman et al., 1996). Direct consumption of brodifacoum baits by Shore Skinks 444 (Oligosoma smithi) in the wild has been observed in New Zealand (Wedding et al., 2010). 445 Wedding et al. (2010) cite records of five other skink species eating cereal baits, some of 446 which contained rodenticides. Bennison et al. 2016 used dye tracers to prove that the large 447 carnivorous King’s Skink (Egernia kingii) had ingested non-toxic baits laid out on islands off 448 the West Australian coast. King’s Skinks were subsequently observed consuming baits 449 containing brodifacoum during the course of a rat eradication on Penguin Island in Western 450 Australia, despite the use of specially designed bait containers intended to exclude the 451 skinks (Bettink, 2015). Others have observed bobtails (Tiliqua rugosa) – another large

452 omnivorous skink – inside AR bait boxes in urban areas (Ashleigh Wolfe, Personal 453 communication).

454 These examples are cause for concern, as both bobtails and large skinks in the genus 455 Egernia have been documented as prey remains at Wedge-tailed Eagle (Aquila audax) nests 456 across a large geographic area (Brooker and Ridpath, 1980). In one instance, remains of 13 457 bobtails were found below a Wedge-tailed Eagle nest on a single visit (Simon Cherriman, 458 unpublished data). Wedge-tailed Eagles are important top carnivores in Australian food 459 webs and are highly susceptible to toxicity from the anticoagulant rodenticide pindone 460 relative to other bird species tested (Martin et al., 1994). Other carnivorous birds and 461 mammals with a higher proportion of reptiles in their diet could potentially be at greater 462 risk.

463 Reptiles could also potentially serve as an effective vector of ARs between 464 invertebrates which consume baits and more sensitive vertebrates at higher trophic levels. 465 Invertebrates have been implicated in directly vectoring rodenticides to bird species 466 including New Zealand Dotterels (Charadrius obscurus aquilonius) (Dowding et al., 2006) and 467 nestling Stewart Island robins (Petroica australis rakiura) (Masuda, Fisher, & Jamieson, 468 2014) as well as the insectivorous European hedgehog (Erinaceus europaeus) (Dowding et 469 al., 2010). If the relative tolerance of ARs demonstrated by Weir et al. (2015) is consistent 470 across numerous reptile taxa, the potential for reptiles to bioaccumulate and biomagnify 471 ARs from lower trophic levels and subsequently retain them for long periods of time makes 472 insectivorous reptiles a potentially important and widely unrecognised vector for 473 anticoagulant rodenticides to more susceptible fauna in higher trophic levels.

474 In Australia, some reptile species, particularly goannas (Varanus spp.), are a 475 culturally and economically important component of a traditional diet for some indigenous 476 peoples (Scelza et al., 2017). Liver tissue of varanids is consumed by some indigenous 477 groups (Caroline Long, Personal communication) and fatty tissues of monitor lizards are 478 eaten preferentially to other body parts (Gracey, 2000). Some rodenticides are known to 479 accumulate to high levels in fat tissue in mammals (Fisher et al., 2015) but accumulation 480 patterns in reptiles are unknown. During the course of a rodent eradication on islands in 481 Western Australia, bungaras (Varanus gouldii) were “observed eating dead and dying rats to

482 the extent that some droppings contained the green dye from the bait” which contained 483 brodifacoum but no mortalities were observed (Burbidge, 2004). These observations raise 484 concerns that if baiting has occurred in or near areas where traditional hunting of varanids 485 takes place, the consumption of varanid tissues likely to accumulate ARs may present a 486 previously unrecognised human health and safety risk. Consumption of feral cats by 487 indigenous people may pose another pathway for rodenticide exposure, as feral cats have 488 been killed by secondary AR poisoning during baiting events in New Zealand (Alterio, 1996). 489 Several studies have cautioned against the consumption of wild game in areas where ARs 490 have been used (Eisemann and Swift, 2006; Pitt et al., 2011), particularly SGARs (Eason et 491 al., 2001). The risks posed by consumption of varanids may be substantially greater than 492 risks associated with rabbit consumption for several reasons. Unlike rabbits which are 493 targeted in discrete baiting events with a FGAR for which there is an established withholding 494 period, varanids are not exposed in a predictable manner and may be chronically exposed to 495 stronger and more persistent SGARs with no established withholding period. The presumed 496 greater physiological tolerance of varanids to ARs and the regular consumption of varanid 497 livers as part of traditional practices considerably elevate the risks associated with varanid 498 consumption relative to rabbit consumption. Urgent investigation of potential rodenticide 499 accumulation in varanids is needed but should take into consideration the high value of this 500 taxon as a traditional food source and the cultural importance of traditional hunting 501 practices. Use of wild reptiles as a food resource is most common in tropical and 502 subtropical areas of the world (Klemens and Thorbjarnarson, 1995) where the prevalence of 503 ARs in wildlife has not been well-studied.

504 The limited literature available suggests that some reptile species are capable of 505 direct bait consumption, long AR retention time, and a capacity to tolerate and biomagnify 506 high concentrations of potent SGARs. These attributes potentially greatly increase the risk 507 of secondary and tertiary vectoring of ARs to more susceptible bird and mammal species in 508 higher trophic levels relative to other regions of the world where small mammals are 509 believed to be the primary vectors. Additional research into the prevalence of AR exposure 510 across a representative sample of reptile taxa will be critical to evaluating the threat of 511 secondary AR poisoning to wildlife in Australia and other countries with high abundance and 512 diversity of reptiles. Depending on the severity and extent of exposure detected, additional

513 work may be warranted to investigate the pathways driving this exposure and the role that 514 reptiles play in vectoring rodenticides to animals in higher trophic levels including humans.

515 Conclusions and Recommendations

516 Most research on exposure of non-target wildlife to ARs has been conducted in cool 517 temperate regions, particularly in North America, Europe, and New Zealand. Patterns of 518 exposure detected in these studies may differ from those in Australia and other tropical and 519 warm arid countries due to differences in the specific ARs used, regulations governing use, 520 and fundamental differences in the taxonomic composition and susceptibility of native 521 fauna. A better understanding of existing knowledge gaps will facilitate more effective and 522 scientifically-informed mitigation measures in Australia and countries with similar climates.

523 In Australia, individuals from 37 species across different feeding guilds, trophic 524 levels, and taxonomic groups have tested positive for AR exposure or are suspected to have 525 been lethally poisoned but most documentation is anecdotal or opportunistic in nature. 526 Instances of poisoning were documented across a wide range of geographic areas but 527 spatial patterns of AR exposure are poorly understood. To date, no thorough investigations 528 directly testing for AR exposure in Australian wildlife have been conducted. Island 529 eradications, feral rabbit control, agricultural application, and residential use have all been 530 implicated as sources of ARs which caused non-target wildlife mortality but the relative 531 contributions of these sources have not been quantified.

532 In aggregate, what little research exists on the interaction between reptiles and ARs, 533 suggests that at least some reptile species may be relatively resistant to the effects but likely 534 to be exposed at high levels. Physiological tolerance, coupled with long retention times 535 could make reptiles effective vectors of ARs in areas of the world where reptiles are 536 abundant. Understanding these dynamics will be critical to understanding the ecology of 537 ARs in tropical and warm arid climates where impacts on wildlife are largely unknown. 538 Effective vectoring of ARs by reptiles poses a potential unevaluated risk to human health in 539 areas where wild reptiles are harvested for human consumption.

540 At present, Australia’s regulatory framework governing the use of ARs is not 541 consistent with emerging practices in other industrialized nations. Restricting SGARs to 542 licensed users and indoor use will likely reduce the incidence and severity of non-target

543 poisoning and the use of lockable bait boxes could reduce risks to children and pets. 544 Coupling these proposed changes with targeted monitoring of rodenticide residues in 545 selected sentinel species will be important in evaluating the efficacy of regulatory changes 546 at reducing non-target mortality. In areas where rodents have developed resistance to 547 FGARs, use of other classes of rodenticides with lower risk of bioaccumulation (such as 548 cholecalciferol) may be a viable option for rodent control with substantially reduced risk of 549 secondary toxicity. At minimum, greater public availability of information on the types, 550 quantities, and locations of ARs sold is necessary to evaluate the risks they pose to non- 551 target wildlife and humans.

552 To address identified knowledge gaps, we suggest the following research priorities:

553  Development of species-specific exposure risk models for carnivorous and 554 omnivorous fauna based on life history parameters

555  Systematic nation-wide testing of multiple taxa of carnivorous and omnivorous 556 wildlife for AR exposure, especially: 557 o species of conservation concern 558 o species consuming small mammals and carrion 559 o marsupial carnivores and scavengers 560 o reptile carnivores and scavengers

561  Systematic long-term testing of geographically widespread and common sentinel 562 species to detect temporal and spatial patterns in AR prevalence

563  Evaluation of the relative contributions of residential, commercial and agricultural 564 use of ARs to wildlife poisoning in Australia 565 o Examine incidence of non-compliance with existing legislation governing AR 566 use 567 o Collection and evaluation of data relating to AR sales and application in 568 Australia

569  Evaluation of the net impact on biodiversity of the use of pindone in and around 570 bushland areas

571  Captive testing of the sensitivity of a wider suite of wildlife species, especially 572 marsupial carnivores and reptiles to SGARs and pindone

573  Examination of the role of reptiles as a vector for ARs in tropical and subtropical 574 nations

575  Evaluation of the risk of rodenticide exposure in humans consuming wild reptiles

576

577 Acknowledgements

578 This project was supported financially by The Holsworth Wildlife Research Endowment via 579 The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird Research Award, 580 the Edith Cowan University School of Science Postgraduate Student Support Award, the 581 Eastern Metropolitan Regional Council’s Healthy Wildlife Healthy Lives program, the Society 582 for the Preservation of Raptors, and Sian Mawson. We thank Allan Burbidge and three 583 anonymous reviewers for improving this manuscript. Images used in the production of the 584 graphical abstract were developed by Tracey Saxby, Jane Hawkey, and Joanna Woerner of 585 the Integration and Application Network, University of Maryland Center for Environmental 586 Science (ian.umces.edu/imagelibrary/).

587

588 Appendix 2.A. Definitions of Schedules applying to all Anticoagulant Rodenticides

589 Registered in Australia from (Australian Government Department of Health:

590 Therapeutic Goods Administration, 2017) 591

592 Schedule 4. – Prescription Only Medicine, or Prescription Animal Remedy – Substances, the 593 use or supply of which should be by or on the order of persons permitted by State or 594 Territory legislation to prescribe and should be available from a pharmacist on prescription.

595 Schedule 5. – Caution – Substances with a low potential for causing harm, the extent of 596 which can be reduced through the use of appropriate packaging with simple warnings and 597 safety directions on the label.

598 Schedule 6. – Poison – Substances with a moderate potential for causing harm, the extent of 599 which can be reduced through the use of distinctive packaging with strong warnings and 600 safety directions on the label.

601 Schedule 7. – Dangerous Poison – Substances with a high potential for causing harm at low 602 exposure and which require special precautions during manufacture, handling or use. These 603 poisons should be available only to specialised or authorised users who have the skills 604 necessary to handle them safely. Special regulations restricting their availability, 605 possession, storage or use may apply.

606 607

608

609 Chapter 3 Anticoagulant rodenticide exposure in an Australian

610 predatory bird increases with proximity to developed habitat

611

612 Lohr, M. T. (2018). Anticoagulant rodenticide exposure in an Australian predatory bird 613 increases with proximity to developed habitat. Science of the Total Environment. 614 643:134–144. https://doi.org/10.1016/j.scitotenv.2018.06.207

615

616 Abstract

617 Anticoagulant rodenticides (ARs) are commonly used worldwide to control 618 commensal rodents. Second generation anticoagulant rodenticides (SGARs) are highly 619 persistent and have the potential to cause secondary poisoning in wildlife. To date no 620 comprehensive assessment has been conducted on AR residues in Australian wildlife. My 621 aim was to measure AR exposure in a common widespread owl species, the Southern 622 Boobook (Ninox boobook) using boobooks found dead or moribund in order to assess the 623 spatial distribution of this potential threat. A high percentage of boobooks were exposed 624 (72.6%) and many showed potentially dangerous levels of AR residue (>0.1mg/kg) in liver 625 tissue (50.7%). Multiple rodenticides were detected in the livers of 38.4% of boobooks 626 tested. Total liver concentration of ARs correlated positively with the proportions of 627 developed areas around points where dead boobooks were recovered and negatively with 628 proportions of agricultural and native land covers. Total AR concentration in livers 629 correlated more closely with land use type at the spatial scale of a boobook’s home range 630 than at smaller or larger spatial scales. Two rodenticides not used by the public 631 (difethialone and flocoumafen) were detected in boobooks indicating that professional use 632 of ARs contributed to secondary exposure. Multiple ARs were also detected in recent 633 fledglings, indicating probable exposure prior to fledging. Taken together, these results 634 suggest that AR exposure poses a serious threat to native predators in Australia, particularly 635 in species using urban and peri-urban areas and species with large home ranges.

636 Introduction

637 Anticoagulant rodenticides (ARs) are commonly used in residential, commercial, and 638 agricultural settings for the control of rodent pests (Rattner et al., 2014b). They block the

639 recycling of vitamin K in the liver, which subsequently disrupts normal blood clotting in 640 vertebrates (Park et al. 1984). ARs are often divided into first generation anticoagulant 641 rodenticides (FGARs) and second generation anticoagulant rodenticides (SGARs) based on 642 their chemical structure and when they were first synthesized. Unlike FGARS, SGARs are 643 often lethal with a single feed and are substantially more persistent in liver tissue (Erickson 644 and Urban, 2004). 645 AR exposure and subsequent mortality have been detected in non-target wildlife in 646 all parts of the world where exposure has been tested (Laakso et al., 2010). Predatory bird 647 species are particularly vulnerable to AR poisoning due to a greater susceptibility to most 648 ARs than other bird species (Herring et al., 2017) and a prey base which frequently contains 649 rodents targeted by the use of ARs. In some raptor species, mortality from AR exposure 650 may have population-level impacts (Thomas et al., 2011). Unlike in Europe and North 651 America, where the non-target impacts of ARs have been extensively studied, relatively 652 little research has been conducted on AR exposure in Australian wildlife (Lohr and Davis, 653 2018; Olsen et al., 2013). This knowledge gap exists despite several lines of evidence 654 suggesting that patterns of regulation and usage in combination with differences in faunal 655 assemblages may increase the incidence and severity of non-target AR poisoning in Australia 656 relative to better-studied areas of the world (Lohr and Davis, 2018). 657 Within Australia, patterns in the spatial distribution of AR exposure have not been 658 studied in any wildlife species. A number of studies have addressed the spatial ecology of 659 anticoagulant rodenticide exposure in non-target wildlife but have been primarily limited to 660 North American mammals. Of these, some have focused on impacts within specific habitat 661 types (Cypher et al., 2014; Gabriel et al., 2012). Studies examining patterns of AR exposure 662 between urban and rural habitats have found correlations between the use of urban habitat 663 and exposure rates in San Joaquin kit foxes (Mcmillin et al., 2008) and bobcats (Riley et al., 664 2007). A model developed to predict exposure patterns in San Joaquin kit foxes found that 665 exposure was most likely in areas of low density housing on the urban/rural interface 666 (Nogeire et al., 2015). Similar dynamics have been suggested but not tested in predatory 667 bird species. Studies in North America and Europe have noted that predatory bird species 668 which use more developed habitats tend to have greater rates of AR exposure than those 669 which predominantly use more natural landscapes (Albert et al., 2010; Christensen et al., 670 2012). Additionally, a study in Spain noted a positive correlation between human

671 population density and AR exposure in a sample of 11 species of predatory birds and 672 mammals (López-Perea et al., 2015). The greater use of rodenticides and higher prevalence 673 of targeted commensal rodents in human-dominated landscapes relative to natural areas is 674 likely to drive these observed and suggested differences in non-target exposure. However, 675 because AR usage patterns differ between urban and agricultural environments (Lohr and 676 Davis, 2018) a need exists to evaluate the possibility of differences in non-target exposure 677 patterns between different types of anthropogenic landscapes. 678 To address this knowledge gap, I sought to compare anticoagulant rodenticide (AR) 679 exposure across intact native bushland and two different types of anthropogenic 680 landscapes. Additionally, I undertook the first large-scale targeted testing of wildlife for AR 681 exposure in the continent of Australia (Lohr and Davis, 2018). Testing was conducted on 682 Southern Boobooks (Ninox boobook), which provide an excellent model to quantify the 683 spatial distribution of threatening processes associated with fragmentation due to their 684 presence across multiple habitat types and high abundance relative to other predatory bird 685 species. To the best of my knowledge, no studies have directly addressed the relative 686 impacts of different types of human land use on AR exposure in non-target wildlife. 687 Understanding how different types of human land use impact the likelihood of AR exposure 688 in non-target wildlife will be critical in evaluating risks to wildlife on a continental scale and 689 will enable more effective targeting of measures to mitigate secondary toxicity.

690 Methods

691 Southern Boobooks are medium-sized hawk owls found across the majority of 692 mainland Australia and adjacent parts of Indonesia and New Guinea (Olsen, 2011a). They 693 are assigned a conservation status of “Least Concern” by the IUCN (“Ninox boobook,” 2018). 694 Some taxonomies consider Southern Boobooks to be synonymous with the closely-related 695 New Zealand Morepork (Ninox novaseelandiae) found in Tasmania and New Zealand but 696 recent genetic and bioacoustic evidence suggests otherwise (Gwee et al., 2017). Boobooks 697 are dietary generalists, consuming a wide variety of vertebrate and invertebrate prey 698 (Higgins, 1999; Trost et al., 2008). These dietary habits make them an ideal model species 699 for broad assessment of contamination of food webs by persistent pollutants like ARs. Their 700 presence in most habitat types across Australia, with the exception of treeless deserts

701 (Higgins, 1999), facilitates examination of differences in exposure across multiple habitat 702 types and allows for future replication of this study at sites across the continent.

703 Specimen Collection 704 Dead boobooks found in Western Australia were solicited from a network of 705 volunteers, wildlife care centres, and government departments and were opportunistically 706 collected when encountered. Boobooks euthanized by veterinarians and wildlife 707 rehabilitators due to severe disease or injury were included. Dates and locations where 708 each boobook was initially collected were recorded from the collector when possible. If 709 liver tissue was identifiable and had a mass >3g, it was removed and stored frozen at 20°C 710 until analysed for AR residues. A total of 73 usable boobook livers were stored for testing. 711 While an effort was made to obtain boobooks from a diversity of geographical areas and 712 habitat types throughout Western Australia, most samples originated in the more densely 713 settled urban and peri-urban areas in the south-west of Western Australia in and around the 714 city of Perth.

715 Rodenticide Analysis 716 Liver samples were analysed by the National Measurement Institute (Melbourne, 717 Australia) for residues of three FGARs (warfarin, coumatetralyl, and pindone) and five SGARs 718 (difenacoum, bromadiolone, brodifacoum, difethialone, and flocoumafen) registered for use 719 in Australia by the Australian Pesticides and Veterinary Medicines Authority. For each 720 sample, 10ml of reverse osmosis water and one gram of liver tissue were added to a 50ml 721 analytical tube and shaken for 15 minutes on a horizontal shaker. A 10ml volume of 5% 722 formic acid in acetonitrile solution was then added and the tube was shaken for an 723 additional 30 minutes. QuEChERS extraction salt was added and the tube was shaken for an 724 additional two minutes. The tube was then centrifuged for 10 minutes at 5100rpm. After 725 pipetting 3ml of the supernatant into a 15 ml analytical tube, 5ml of hexane was added and 726 the tube was shaken for two minutes then centrifuged for 10mins at 5100rpm. The hexane 727 layer was removed using a vacuum pipette and discarded. A 1ml aliquot of the supernatant 728 was transferred to a 2ml QuEChERS dispersive tube, shaken for one minute, and centrifuged 729 at 13000rpm for three minutes. The QuEChERS supernatant was then filtered using a 730 0.45μm filter. After filtration, 3μl of coumachlor was added as an internal standard to 497μl 731 of the filtered extract and vortexed prior to LC-MS/MS analysis. A Waters TQS Tandem

732 Quadrupole Detector Liquid Chromatograph-Mass Spectrometer (LC-MS/MS) and an 733 Acquity UPLC CSH C18 100 x 2.1mm column were used to quantify concentrations of each 734 rodenticide. Recovery rates for each AR, were calculated using chicken liver samples spiked 735 with analytical standards (Table 3.1). 736 Table 3.1 Limit of detection (LOD), limit of quantification (LOQ), average recovery, and relative standard deviation (RSD) for 737 eight ARs in a spiked chicken liver matrix.

Compound LOD (mg/kg) LOQ (mg/kg) Average recovery % (RSD) Warfarin 0.001 0.002 94 (8.1) Coumatetralyl 0.001 0.002 93 (7.6) Bromadiolone 0.005 0.010 96 (9.5) Difenacoum 0.005 0.010 96 (11.2) Flocoumafen 0.005 0.010 103 (11.4) Brodifacoum 0.005 0.010 92 (8.8) Difethialone 0.005 0.010 91 (14.6) Pindone 0.005 0.010 36 (13.5) 738

739 Statistical Analysis 740 Total AR liver concentration is commonly used to compare toxicity risk when 741 individuals are exposed to multiple rodenticides (Christensen et al., 2012) due to similarities 742 in their modes of action and likely cumulative effects (Hughes et al., 2013). For this reason, 743 the sum of all liver rodenticide concentrations above the limit of detection was calculated 744 for each individual for the purposes of comparing differences in exposure by age, season, 745 and land use. In order to compare seasonal trends in total AR concentration, boobooks 746 were assigned to four groups based on their collection date: summer (December – 747 February), autumn (March-May), winter (June- August), and spring (September-November). 748 All boobooks with known collection months (n=71) were included in the seasonal analysis. 749 The Kruskal-Wallis test was used to assess whether significant differences existed in liver AR 750 concentration by season. 751 Boobooks were assigned to age classes of less than one year ("hatch year") or 752 greater than one year ("after hatch year") based on the presence of juvenile down and by 753 examination of fluorescence patterns under ultraviolet light (Weidensaul et al., 2011). In 754 one instance, it was not possible to determine age class due to degradation of porphyrins 755 caused by prolonged exposure of ventral remiges to sunlight. A total of 72 boobooks of 756 determined age class were available for analysis of the relationship between age and AR 757 exposure. I used a Mann-Whitney-Wilcoxon test to determine whether total liver

758 concentration of ARs varied between the two age classes. Results were considered 759 significant if p<0.05.

760 Exposure Thresholds 761 The utility of rodenticide concentration in liver tissue as a means to diagnose lethal 762 exposure has been questioned (Erickson and Urban, 2004; Thomas et al., 2011) as 763 susceptibility to acute toxicity can vary among individuals and across species (Thomas et al., 764 2011). Exposure to multiple ARs adds additional complexity to the assessment of likely 765 impacts from residual liver concentrations (Murray, 2017). However, a need exists to 766 estimate likely impacts across exposed individuals and to compare the magnitude of 767 exposure to previous studies. Accordingly, I identified relevant literature which established 768 commonly used guidelines for outcomes of various exposure rates in related taxa to allow 769 estimation of likely impacts on boobooks. 770 The Rodenticide Registrants Task Force suggested that a 0.7 mg/kg liver 771 concentration of brodifacoum was likely to be toxic based largely on captive studies of Barn 772 Owls (Kaukeinen et al., 2000), however this threshold estimate may be too high, as 773 environmental conditions affecting wild birds may increase their susceptibility to ARs 774 relative to captive birds (Mendenhall and Pank, 1980). Dowding et al. (1999) estimated a 775 lethal liver concentration for brodifacoum of 0.5 mg/kg using 29 individuals from 10 species 776 of birds. Numerous studies have reported thresholds of 0.2 mg/kg (Albert et al., 2010; 777 Christensen et al., 2012; Hughes et al., 2013; Langford et al., 2013; López-Perea et al., 2015; 778 Stansley et al., 2014; Walker et al., 2008) and 0.1 mg/kg (Albert et al., 2010; Christensen et 779 al., 2012; Langford et al., 2013; Ruiz-Suárez et al., 2014; Shore et al., 2016; Stansley et al., 780 2014; Walker et al., 2011, 2008) as indices of lower limits at which lethal AR toxicity was 781 likely to occur in predatory birds. These estimates were based on two studies examining 782 wild barn owls: Newton et al. (1999) and Newton et al. (1998) respectively. I also included a 783 threshold of 0.01mg/kg as this is the lowest published record of lethal SGAR toxicity in a 784 predatory bird species (Stone et al., 1999). Boobook liver concentrations were compared 785 against these thresholds (0.7 mg/kg, 0.5mg/kg, 0.2 mg/kg, 0.1 mg/kg, and 0.01mg/kg) to 786 facilitate a comprehensive understanding of overall potential impacts of ARs across all 787 sampled individuals.

788 Spatial Analysis 789 Only boobooks with accurate location data were included in the spatial analysis. In 790 one instance, two road-killed boobooks were recovered at the same location. One of these 791 was randomly removed from the spatial analysis, leaving a total of 66 boobooks available 792 for analysis. Land cover for the state of WA was classified into developed, agriculture, 793 native vegetation or open water. The developed category included all areas with 794 anthropogenic impervious surfaces (roads, buildings car parks, etc.) as well as intensive land 795 uses that did not qualify as agriculture (mines, landfills, sports grounds, golf courses etc.). 796 The agriculture category included a diversity of irrigated and dryland crops, orchards, and 797 grazed areas. Intensive indoor animal agriculture was included in the developed category 798 rather than agriculture because it consisted primarily of buildings and other impervious 799 surfaces. Areas subjected to silvicultural practices were classified as part of the native 800 vegetation category due to structural similarity. Additionally like native bushland, the only 801 anticoagulant permitted for use in forestry is pindone which is used to control rabbits in 802 areas too close to human habitation to allow the safe use of 1080. Percentages of each 803 classification were calculated within circular buffer zones (areas of influence) of three 804 different sizes around each location where a boobook was found. The two smaller buffer 805 sizes were calculated to match the mean area of a boobook’s core home range (7.3 ha) and 806 total home range (145.1 ha) (Olsen et al., 2011). The largest buffer size was an arbitrarily 807 large area with a 3km radius (2827.4 ha). This larger buffer was included to account for the 808 possibility of movement of contaminated prey into boobooks’ home ranges from adjacent 809 areas influencing the probability of boobook exposure to ARs. Because open water was not 810 considered to be usable space, the percentages of the other three habitat types were 811 calculated excluding any open water within the buffers. 812 I used general linear models with a negative binomial distribution, following 813 methodology used by Christensen et al. (2012), to analyse differences in rodenticide 814 exposure by habitat composition at the three different spatial scales. The Akaike 815 Information Criterion AIC was used to rank models for habitat proportions at each spatial 816 scale. Only single variable models were considered in the ranking due to nesting and 817 correlation of habitat proportions and spatial scales. I calculated McFadden's pseudo-R2 818 values for each habitat type and spatial scale combination. Statistical analysis was 819 performed using RStudio 1.1.383 (RStudio, Inc., Boston, MA, USA).

820 Results

821 While I did not directly quantify physiological signs of rodenticide poisoning due to 822 most carcasses being damaged as a result of vehicle collisions, during dissection I observed 823 symptoms associated with acute lethal AR toxicity in at least nine boobooks exhibiting no 824 sign of trauma. These symptoms included excessive bleeding from minor lacerations, pale 825 or mottled livers, subdermal and muscular haemorrhage in the absence of trauma, blood in 826 the thoracic cavity, and blood around the mouth and nares. Similar symptoms have been 827 described in association with lethal AR toxicity in other raptor species (Murray, 2017).

828 Table 3.2 Percentage exposure, mean exposure and total detection of eight different anticoagulant rodenticides in livers of 73 Southern Boobooks in Western Australia.

Coumatetralyl Warfarin Pindone Difenacoum Brodifacoum Bromadiolone Difethialone Flocoumafen Total Percent Exposed 0.000 2.740 0.000 15.068 72.603 31.507 8.219 2.740 72.603 Mean Exposure (mg/kg) 0.000 0.000 0.000 0.004 0.260 0.019 0.015 0.011 0.310 Standard Error 0.000 0.000 0.000 0.002 0.064 0.005 0.011 0.011 0.069 Maximum Concentration (mg/kg) 0.000 0.002 0.000 0.097 4.002 0.214 0.775 0.818 4.002 Minimum Concentration (mg/kg) 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 Total Detected (mg/kg) 0.000 0.003 0.000 0.287 18.994 1.421 1.063 0.834 22.606 829

830

831

832

833

834

835

836

837

838

839

840 Table 3.3 Published rates of multiple second generation anticoagulant rodenticide exposure and percentages of individuals with exposure above two thresholds in predatory birds.

% Multiple % >0.1 % >0.2 Mean Exposure Species Location n Individuals % Exposed Source Exposure mg/kg mg/kg (mg/kg) (SE)

Southern Boobook (Ninox boobook) Western Australia 73 72.6 38.4 50.7 35.6 0.310 (0.069) this study Tawny Owl (Strix aluco) United Kingdom 172 19.2 2.9 12.2 5.8 0.125 Walker et al., 2008 Barn Owl (Tyto alba) United Kingdom 94 72 16 Shore et al., 2016 100 Red Kite (Milvus milvus) Scotland 69.3 36 17.5 0.155 (0.017) Hughes et al., 2013 114 Buzzard (Buteo buteo) Scotland 44.3 14.2 2.1 0.047 (0.004) Hughes et al., 2013 479 Kestrel (Falco tinnunculus) Scotland 40.9 17.4 9.1 0.173 (0.082) Hughes et al., 2013 22 Barn Owl (Tyto alba) Scotland 34.9 17.5 17.5 0.076 (0.018) Hughes et al., 2013 63 Tawny Owl (Strix aluco) Scotland 38.2 5.9 2.9 0.047 (0.021) Hughes et al., 2013 34 Sparrowhawk (Accipiter nisus) Scotland 37 54.1 29.7 2.7 0.060 (0.016) Hughes et al., 2013 Peregrine Falcon (Falco peregrinus) Scotland 24 29.2 0 0 0.017 (0.007) Hughes et al., 2013 Barn Owl (Tyto alba) United Kingdom 52 Walker et al., 2011 58 84 17.2 Red Kite (Milvus milvus) United Kingdom 89 Walker et al., 2011 18 94 Kestrel (Falco tinnunculus) United Kingdom 95 Walker et al., 2011 20 100 Barn Owl (Tyto alba), Barred Owl (Strix varia), and Great Horned Owl (Bubo Canada 15 Albert et al., 2010

virginianus) 164 92 32 0.107 Great Horned Owl Canada 123 0.016 Thomas et al., 2011

Red-tailed Hawk (Buteo jamaicensis) Canada 58 0.005 Thomas et al., 2011

Golden eagle (Aquila chrysaetos) Norway 16 73.3 31.3 25 6.3 0.051 Langford et al., 2011 Eagle owl (Bubo bubo) Norway 8 62.5 25 37.5 12.5 0.087 Langford et al., 2011 Osprey (Pandion haliaetus) Norway 3 0 0 0 0 0 Langford et al., 2011 Peregrine falcon (Falco peregrinus) Norway 2 0 0 0 0 0 Langford et al., 2011 Gryfalcon (Falco rusticolus) Norway 1 0 0 0 0 0 Langford et al., 2011 Red-tailed Hawk (Buteo jamaicensis) USA 37 97 78 Murray, 2017

Barred Owl (Strix varia) USA 24 88 42 Murray, 2017

Great Horned Owl (Bubo virginianus) USA 17 100 71 Murray, 2017

Eastern Screech-Owl (Megascops asio) USA 16 100 69 Murray, 2017

Red-tailed Hawk (Buteo jamaicensis) USA 105 81 15 47 25 0.117 Stansley et al., 2014 Great Horned Owl (Bubo virginianus) USA 22 82 18 36 9 0.07 Stansley et al., 2014 Eurasian Sparrowhawk (Accipiter nisus) Spain (Canary Islands) 14 85.7 0.0577 Ruiz-Suárez et al., 2014

Long-eared Owl (Asio otus) Spain (Canary Islands) 23 73.9 0.1322 Ruiz-Suárez et al., 2014

Common Buzzard (Buteo buteo) Spain (Canary Islands) 9 26.3 0.0368 Ruiz-Suárez et al., 2014 Spain (Canary Islands) 0.0915 Ruiz-Suárez et al., 2014 Barbary Falcon (Falco pelegrinoides) 16 31.2 Kestrel (Falco tinnunculus) Spain (Canary Islands) 21 66.6 0.219 Ruiz-Suárez et al., 2014

Barn Owl (Tyto alba) Spain (Canary Islands) 21 76.2 0.1344 Ruiz-Suárez et al., 2014

All Species Spain (Canary Islands) 104 63.5 34.8 Ruiz-Suárez et al., 2014

Scops Owl (Otus scops) Spain (Majorca Island) 26 57.7 0 0.0134 López-Perea et al., 2015

Barn Owl (Tyto alba) Spain (Majorca Island) 19 84.2 57.9 0.2337 López-Perea et al., 2015

Scops Owl (Otus scops) Spain (Catalonia) 7 14.3 0 0.1584 López-Perea et al., 2015

Barn Owl (Tyto alba) Spain (Catalonia) 22 54.5 13.6 0.1178 López-Perea et al., 2015

Tawny Owl (Strix aluco) Spain (Catalonia) 27 77.8 29.6 0.0952 López-Perea et al., 2015

Eagle Owl (Bubo bubo) Spain (Catalonia) 14 100 64.3 0.2896 López-Perea et al., 2015

Long-eared Owl (Asio otus) Spain (Catalonia) 12 58.3 0 0.0111 López-Perea et al., 2015

Little Owl (Athene noctua) Spain (Catalonia) 7 71.4 28.6 0.1972 López-Perea et al., 2015

Common buzzard (Buteo buteo) Spain (Catalonia) 56 64.3 26.8 0.1253 López-Perea et al., 2015

Barn owl (Tyto alba) Denmark 80 94 37.4 13.7 0.1141 Christensen et al., 2012 Buzzard (Buteo buteo) Denmark 141 94 20.6 5.7 0.0745 Christensen et al., 2012 Eagle owl (Bubo bubo) Denmark 10 100 70 70 0.1931 Christensen et al., 2012 Kestrel (Falco tinnunculus) Denmark 66 89 27.2 13.6 0.099 Christensen et al., 2012 Little owl (Athene noctua) Denmark 9 100 33.3 22.2 0.1186 Christensen et al., 2012 Long-eared owl (Asio otus) Denmark 38 95 0 0 0.0194 Christensen et al., 2012 Marsh harrier (Circus aeruginosus) Denmark 3 100 0 0 0.0123 Christensen et al., 2012 Red kite (Milvus milvus) Denmark 3 100 0 66.7 0.413 Christensen et al., 2012 Rough-legged Buzzard (Buteo lagopus) Denmark 31 84 12.9 0 0.0408 Christensen et al., 2012 Short-eared owl (Asio flammeus) Denmark 5 100 0 0 0.015 Christensen et al., 2012 Tawny owl (Strix aluco) Denmark 44 93 20.5 9.1 0.0784 Christensen et al., 2012

All Species Denmark 430 73 Christensen et al., 2012

841 ARs were detected in 72.6% of all boobook liver samples (Table 3.2) with a mean 842 summed AR exposure of 0.310 mg/kg (SE 0.069246735) (Table 3.3). Approximately 17.8% of 843 boobook livers contained greater than the suspected lethal threshold of 0.5 mg/kg total ARs 844 (Figure 3.1) with 13.7% above the more conservative limit of 0.7 mg/kg. Seven of the ten 845 boobooks with AR liver concentrations above 0.7 mg/kg appear to have died directly of AR 846 poisoning and the other three showed signs of poisoning described by Murray (2017) 847 despite other apparent proximate causes of death. More than half of the boobooks tested 848 had liver concentrations above 0.1 mg/kg (Figure 3.1) and would likely have experienced at 849 least some degree of coagulopathy (Rattner et al., 2014a). The majority of boobooks 850 (65.8%) were exposed at a level above 0.01 mg/kg – the lowest observed lethal threshold in 851 an owl (Figure 3.1).

852

853 Figure 3.1 Percentages of Southern Boobooks (n=73) in Western Australia exposed to rodenticides stratified by total 854 rodenticide liver concentration (mg/kg) thresholds indicating potential outcomes.

855 The three FGARs tested – coumatetralyl, warfarin, and pindone – were infrequently 856 detected and accounted for only 0.01% of all ARs detected (Table 3.2). Coumatetralyl and 857 pindone were not detected in any of the samples and warfarin was detected in two 858 individuals at low levels (0.0024 mg/Kg and 0.0014 mg/Kg). The lower of these was below 859 the limit of quantification. Detectable exposure to SGARs was substantially higher (Table 860 3.2). Brodifacoum – the most commonly detected SGAR – was found in 72.6% of samples

861 and made up 84.0% of all rodenticides detected by mg/kg. It was detected in all liver 862 samples containing AR residues (Table 3.2). Difethialone and flocoumafen, which were not 863 known to be in use by the public were also detected in boobooks. Two or more ARs were 864 detected in 38.4% of boobooks tested (Figure 3.2). A maximum of five different ARs was 865 detected in two individual boobooks.

866

867 Figure 3.2 Percentages of Southern Boobooks (n = 73) exposed to multiple anticoagulant rodenticides in Western Australia.

868 Mean total liver concentration of ARs was not significantly different between age 869 classes (p= 0.34). AR exposure was greatest in boobooks collected in winter and winter 870 concentrations were significantly different from summer concentrations (p=0.026) (Figure 871 3.3). The livers of two recent fledglings still under parental care contained low but 872 quantifiable amounts of brodifacoum (0.022 and 0.051 mg/kg) and difethialone (0.020 and 873 0.022 mg/kg).

874

875 Figure 3.3 Mean total anticoagulant rodenticide concentration (mg/kg) in liver tissue of Southern Boobooks (n= 71) in 876 Western Australia by season.

877 Total AR exposure was positively correlated with the amount of developed area 878 within buffers at all spatial scales (Table 3.4). Proportions of agriculture and bushland 879 habitat within buffers were negatively correlated with total AR exposure at all spatial scales 880 (Table 3.4). The three AIC top-ranked models quantified habitat composition at the scale of 881 a full boobook home range and were all statistically significant (Table 3.4). The top-ranked 882 model used developed habitat at the scale of a boobook’s total home range and was highly 883 significant (p=0.00182). Correlations between the top three ranked models and total AR

884 concentration were not particularly strong but are stronger than would be suggested by 885 interpretation of traditional R2 indices, as McFadden's pseudo-R2 values falling in the range 886 of 0.2 to 0.4 “represent an excellent fit” (McFadden, 1978). 887 Table 3.4 Akaike information criterion (AIC) ranking of models of the association between percentage of single land use 888 types within buffers around collection points and total anticoagulant rodenticide liver concentration in Southern Boobooks 889 (n= 66) in Western Australia at three different spatial scales (Big=2827.4 ha buffer, Mid=145.1 ha buffer, Small=7.3 ha 890 buffer.

Std. Model Estimate Error z value Pr(>|z|) AIC McFadden's pseudo-R2 Mid Developed 2.1439 0.6876 3.118 0.00182 751.43 0.08675021 Mid Agriculture -2.4505 0.9844 -2.489 0.0128 754.28 0.05158204 Mid Native 0.05081192 Vegetation -2.5139 0.9584 -2.623 0.00871 754.35 Big Agriculture -3.0121 1.1147 -2.702 0.00689 754.51 0.04870524 Small Developed 1.5092 0.6822 2.212 0.027 754.53 0.04854103 Big Developed 1.7553 0.7547 2.326 0.02 754.83 0.04473145 Small Agriculture -1.6016 1.0237 -1.565 0.118 756.27 0.02641717 Small Native Vegetation -1.364 0.9249 -1.475 0.14 756.59 0.02232542 Big Native 0.01968855 Vegetation -1.9017 1.066 -1.784 0.0744 756.8 891

892 Discussion

893 The overall proportion of boobooks with detectable AR exposure (72.6 %) and the 894 proportion of boobooks exposed to two or more rodenticides (38.4%) was high but within 895 the range of estimates generated by studies in Europe and North America (Table 3.3). Mean 896 total AR concentration in boobooks (0.310 mg/kg) was substantially higher than any other 897 available published estimate with the exception of Red Kites (Milvus milvus) (0.413 mg/kg) 898 in Denmark (Christensen et al., 2012). The extremely high mean exposure in boobooks may 899 result from multiple causes. A large proportion of samples were obtained from urban and 900 peri-urban areas where exposure is likely to be more prevalent. This was also the case in 901 several other studies documenting high exposure rates and liver concentrations (López- 902 Perea et al., 2015; Murray, 2017; Stansley et al., 2014). As a consequence, the sample of 903 boobooks used in this study is probably not representative of Australia as a whole but may 904 provide a useful estimate for other large human population centres elsewhere. Circadian 905 activity patterns may also increase boobooks’ risk of AR exposure relative to some other 906 raptor species. Nocturnal species have been noted to have higher liver AR concentrations

907 than diurnal species (Ruiz-Suárez et al., 2014; Sánchez-Barbudo et al., 2012). If owls using 908 highly populated landscapes are at greater risk than other bird species, future evaluation of 909 Powerful Owls which use urban and peri-urban areas and are listed as vulnerable in Victoria 910 may be warranted. Southwest populations of Masked Owls (Tyto novaehollandiae) and 911 Barking Owls (Ninox connivens), both of which are listed as P3 priority fauna (poorly known 912 but thought to be possibly threatened) in Western Australia, may also be susceptible to AR 913 poisoning in areas where developed habitats are encroaching on their remaining ranges. 914 As a consequence of the methodology used in sample collection, this study probably 915 underestimates the proportion of lethal poisonings which actually occur. Anticoagulant 916 rodenticides induce lethargy prior to mortality and lethally poisoned owls are more likely to 917 die in nest hollows or roost sites in dense vegetation where their likelihood of detection by 918 humans would be low (Newton et al., 1990). Similar underestimation of lethal toxicity has 919 been suggested in studies of mammals exposed to ARs, as well (Mcdonald et al., 1998). 920 Conversely, if haemorrhaging induced by sub-lethal exposure reduced a boobook’s reaction 921 time or ability to fly, it could increase the risk of other proximate sources of mortality 922 (Newton et al., 1990) such as collisions with vehicles or windows. This could potentially 923 increase its likelihood of being killed in a conspicuous location and subsequently collected 924 for this study with the end result of inflating the number of sub-lethally exposed birds 925 entering this study.

926 Individual Rodenticides 927 A lack of detectable pindone residues in the livers of the boobooks sampled was 928 unexpected because pindone is used within the Perth metropolitan area to control rabbits 929 in urban bushlands and previous literature implicates similar control programs elsewhere in 930 Australia in secondary poisonings of native raptors (Olsen et al., 2013) though this has 931 recently been disputed (Olsen and Rae, 2017). Failure to detect pindone could be the result 932 of a short retention time relative to more persistent SGARs (Fisher et al., 2003), its use in 933 targeted and short-term control efforts, low overall usage relative to commercial and 934 residential use of other anticoagulant rodenticides, or dietary patterns of boobooks 935 precluding consumption of European rabbits (Oryctolagus cuniculus) – the species targeted 936 by pindone applications. While it is possible that occasional localised exposure may occur, it 937 appears that pindone, as currently applied in urban and peri-urban areas does not 938 constitute a substantial threat to boobook populations relative to other rodenticides

939 originating from commercial and residential sources. Future studies on impacts of pindone 940 on native raptors should consider testing species which are more likely to prey on rabbits 941 (Wedge-tailed Eagles (Aquila audax) and Little Eagles (Hieraaetus morphnoides)) (Olsen et 942 al., 2006) or scavenge rabbit carcasses (Whistling Kites (Haliastur sphenurus)) (Fuentes et al., 943 2005) and are at greater risk of secondary exposure. 944 Failure to detect coumatetralyl in any samples and the detection of warfarin at 945 extremely low concentration in only two samples despite commercial availability to the 946 public suggests that their relatively short half-life in liver tissue (Fisher et al., 2003) probably 947 reduces the incidence and severity of secondary exposure and precludes bioaccumulation 948 and biomagnification. This result is consistent with absence or low concentration and 949 prevalence of FGARs relative to SGARs in other wildlife species since SGARs came into 950 widespread use (Albert et al., 2010; Fourel et al., 2018; Murray, 2017; Ruiz-Suárez et al., 951 2014). 952 The detection of brodifacoum at rates an order of magnitude higher than all other 953 ARs combined is probably attributable to a combination of its greater duration of 954 persistence in liver tissue (Horak et al., 2018), more prevalent use, and incorporation into a 955 greater number of commercially available rodenticide bait products. This is particularly 956 concerning because captive studies suggest that brodifacoum is more likely to cause 957 secondary toxicity in birds than any other tested ARs due to its high toxicity and long liver 958 retention time (Erickson and Urban, 2004). Bromadiolone and difenacoum respectively, 959 were the next most commonly detected in samples (Table 3.2). This is probably because, 960 together with brodifacoum, they comprise the three SGARs commonly available in WA at 961 retail stores. At present, brodifacoum, bromadiolone, and difenacoum probably pose the 962 greatest threat of secondary poisoning to non-target wildlife of all ARs in use. 963 The detection of flocoumafen and difethialone – which are not readily available to 964 the public due to sale in bulk quantities but are used by pest control professionals – 965 indicates that at least some proportion of wildlife exposure is directly related to commercial 966 pest control activities. Flocoumafen was the most prevalent rodenticide detected in liver 967 tissue of one boobook, which died shortly after admission to a wildlife care centre and 968 showed physiological signs of AR poisoning (pale mottled liver, subcutaneous haemorrhage, 969 and large quantities of blood in the abdominal cavity). These findings have potentially 970 serious implications for legislation attempting to curtail non-target exposure by limiting

971 public access to SGARs. In the United States, legislation restricting the use of SGARs to 972 licensed professionals went into effect in 2011 (Bradbury, 2008). However, a subsequent 973 study found an increase in AR exposure in four predatory bird species in Massachusetts, 974 USA following the ban (86% of 161 birds from 2006 - 2010 compared to 96% of 94 birds 975 exposed from 2012 - 2106) perhaps due to an increased use of professional rodent control 976 services (Murray, 2017). My findings provide additional evidence that use of ARs by 977 professional pesticide applicators does contribute, at least to some degree, to poisoning of 978 non-target raptors. However, the impacts of this source relative to private use are difficult 979 to assess because other SGARs which are available to the public – particularly brodifacoum 980 and bromadiolone – are in common use by professional pesticide applicators in WA. Taken 981 together, these results cast doubt on whether regulations restricting sale of SGARs from 982 private use will be sufficient to reduce widespread exposure and toxicity in predatory birds. 983 After the completion of this study, it was brought to my attention that diphacinone 984 was also being used in Western Australia by commercial pesticide applicators. This FGAR 985 has a relatively short half-life of three days in rat liver tissue and as a consequence is 986 unlikely to bioaccumulate and cause secondary poisoning in predatory non-target wildlife 987 (Fisher et al., 2003). The registration of diphacinone in Australia has expired. However, if 988 diphacinone is re-registered, future monitoring projects should include diphacinone testing 989 as it could potentially contribute to overall rodenticide exposure. 990 Exposure to multiple rodenticides (38.4%) was relatively common in sampled 991 boobooks but not as frequent as in some other predatory bird species (Christensen et al., 992 2012; Murray, 2017; Walker et al., 2011). The relatively high rate of multiple exposure and 993 the presence of detectable levels of up to five different ARs in liver tissue suggests 994 cumulative exposure from multiple prey items over an extended period of time. This 995 hypothesis is supported by the finding that livers of adult raptors in Denmark contained 996 multiple rodenticides more frequently than those of juveniles (Christensen et al., 2012). The 997 prevalence of multiple exposures in boobooks is particularly concerning because laboratory 998 studies on rats determined that warfarin sensitivity is increased after sub-lethal exposure to 999 brodifacoum (Mosterd and Thijssen, 1991). If ARs have a synergistic effect rather than a 1000 purely additive effect, raptors may be negatively impacted at a lower threshold when 1001 exposed to more than one AR, leading to underestimates of negative impacts on non-target 1002 wildlife.

1003 Rodenticide Thresholds 1004 The utility of detectable rodenticide concentration in liver tissue as a means to 1005 diagnose lethal exposure has been questioned (Erickson and Urban, 2004; Thomas et al., 1006 2011) as susceptibility to acute toxicity can vary among individuals and across species 1007 (Erickson and Urban, 2004). However, it can be informative in comparing environmental 1008 exposure and as an index for potential impacts at the population level. Depending on the 1009 threshold used (0.7 mg/kg or 0.5 mg/kg), either 13.7% or 17.8% of boobooks tested had 1010 rates of exposure consistent with likely lethal outcomes. Confirmation of physical signs of 1011 rodenticide poisoning in all boobooks with AR liver concentrations above 0.7 mg/kg and the 1012 absence of other obvious causes of death in 70% of these individuals indicates that this 1013 threshold is a reasonable guideline for estimating likely lethal toxicity in boobooks. 1014 Regardless of the threshold used, the relatively high frequency of exposure at levels likely to 1015 be directly lethal is cause for concern. In combination with visible signs of AR poisoning, it 1016 indicates that exposure to ARs contributed substantially to mortality in bobooks found dead 1017 or brought to wildlife carers in the urban and peri-urban areas where most samples were 1018 collected. 1019 Exposure at potentially dangerous but not necessarily lethal levels was also high 1020 relative to most published studies examining rodenticide exposure in wild raptors found 1021 dead or moribund. The proportion of boobooks exposed at levels above 0.2 mg/kg (35.6%) 1022 was higher than all other reported estimates except for in Barn Owls (Tyto alba) (57.9%) and 1023 Eagle Owls (Bubo bubo) (64.3%) in Spain (López-Perea et al., 2015) and Red Kites (Milvus 1024 milvus) (66.7%) in Denmark (Christensen et al., 2012). In all three species, the sample size 1025 was small (n<20). The percentage of boobooks with total AR liver concentrations above 1026 0.1mg/kg (50.7%) was substantially greater than all previously reported species except for 1027 Red-tailed Hawks in New Jersey, USA (47%) (Stansley et al., 2014). At minimum, a 1028 threshold of 0.1 mg/kg should be considered potentially dangerous. In a laboratory study 1029 using Eastern Screech Owls (Megascops asio), diphacinone concentrations of ≥0.1 mg/kg in 1030 liver tissue were associated with coagulopathy (Rattner et al., 2014a). Coagulopathy is likely 1031 more dangerous to wild birds due to greater amounts of movement and injuries associated 1032 with capturing prey and may have synergistic interactions with environmental stressors 1033 which increase the chance of mortality (Erickson and Urban, 2004). SGARs are also more

1034 toxic than diphacinone and can logically be expected to have at least as great of an impact 1035 at the same threshold. 1036 Sub-lethal exposure was common in boobooks regardless of the chosen threshold. 1037 The sub-lethal impacts of chronic AR exposure are poorly studied in wildlife. A number of 1038 lines of evidence suggest that even exposure below the threshold needed to cause lethal 1039 haemorrhage is not benign. While Thomas et al. (2011) take issue with the uncritical use of 1040 liver concentrations to assess likely toxicity, their probabilistic methodology examining AR 1041 toxicity in four raptor species predicted that 20% of individuals would experience 1042 quantifiable toxicity at levels as low as 0.08 mg/kg. Increased rates of parasitism and 1043 infectious disease have also been documented in association with AR exposure in bobcats 1044 (Lynx rufus) (Riley et al., 2007), Great Bustards (Otis tarda) (Lemus et al., 2011), and 1045 common voles (Microtus arvalis) (Vidal et al., 2009). In bobcats, immunosuppression and 1046 inflammatory response associated with chronic sub-lethal AR exposure and use of urban 1047 habitats may have led to an outbreak of notoedric mange (Serieys et al., 2018). Similar 1048 disruption of immune system function may occur in other chronically-exposed wildlife 1049 (Serieys et al., 2018). Several studies have also suggested the possibility of increased 1050 mortality rates via accidents, predation, vehicle collisions, nutritional stress, and blood loss 1051 following minor injury in wildlife exposed to sub-lethal doses of anticoagulant rodenticides 1052 (Albert et al., 2010; Mendenhall and Pank, 1980; Newton et al., 1990; Stone et al., 2003, 1053 1999). If this dynamic is indeed consistent across wildlife species, the high rates of 1054 presumably sub-lethal exposure detected in boobooks are cause for concern. If sub-lethal 1055 exposure to ARs substantially increases the risk of parasitism and other sources of mortality, 1056 it is not appropriate to assess the overall impacts of anticoagulants on predatory bird 1057 populations based solely on documentation of direct lethal toxicity.

1058 Spatial Correlations 1059 We observed weak but statistically significant correlations between AR exposure and 1060 habitat proportions in proximity to recovered boobook carcasses. The difference in the 1061 direction of correlations between AR exposure and proportions of agricultural and 1062 developed habitats, the consistency of the trends at different spatial scales, and the 1063 increasing strength of the trends at the most biologically meaningful spatial scale all suggest 1064 an actual difference in exposure risk between the two anthropogenic landscapes. Future 1065 studies on this topic should attempt to improve sample collection across different types of

1066 anthropogenic landscapes or focus on species for which samples are more readily available 1067 across study areas. A low sample size of boobook carcasses from landscapes predominantly 1068 comprised of native bushland or agriculture likely contributed to the low predictive value of 1069 top models. 1070 The three top-ranked models for boobook AR exposure used habitat data at the 1071 scale of an average home range. Foraging behaviour likely explains the closer correlation of 1072 AR exposure and habitat type at the spatial scale of an average boobook home range 1073 relative to other spatial scales. The vast majority of foraging occurs within an animal’s home 1074 range and its exposure to ARs can be expected to relate most closely to the proportions of 1075 habitat types likely to be sources of contamination of its prey base at this spatial scale. 1076 Boobooks have relatively small home ranges in comparison to other Australian owl species 1077 (Kavanagh and Murray, 1996; Soderquist and Gibbons, 2007). If risk of rodenticide exposure 1078 is related to developed area at the scale of an animal’s home range, species with larger 1079 home ranges may be exposed over a broader portion of the landscape. This hypothesis is 1080 supported by the finding that in bobcats – a species with a much larger home range than 1081 boobooks– the concentration but not the presence of ARs in liver tissue correlated with the 1082 proportion of developed habitat within their home range (Riley et al., 2007). Taken in 1083 combination, these results suggest that species with large home ranges are likely to be at 1084 risk of some degree of AR exposure if their home range encompasses even small areas of 1085 developed habitat. As a consequence, encroachment of human structures into large areas 1086 of natural habitat may have an impact on predatory species with large home ranges that is 1087 disproportionate to the area of habitat lost through development. 1088 The positive correlation between total AR exposure and the proportion of developed 1089 area within buffers was expected due to the widespread use of rodenticides in commercial 1090 and residential settings. This pattern of exposure has been suggested following detection of 1091 high exposure rates in densely populated areas (López-Perea et al., 2015; Stansley et al., 1092 2014) but, this appears to be the first instance where differences in exposure across habitat 1093 types has been directly quantified in a bird species. A number of other studies have 1094 examined the spatial patterns of AR exposure in wildlife. The trend in boobooks was similar 1095 to the correlation between developed areas and total AR exposure observed in a study of 1096 bobcats and mountain lions in California (Riley et al., 2007). Similarly, AR exposure was 1097 common (87%) in an urban population of San Joaquin kit foxes but no rodenticides were

1098 detected in individuals from a non-urban population (Mcmillin et al., 2008). Frequent AR 1099 exposure in wildlife inhabiting developed habitats is typically attributed to the “prevalent 1100 and wide-spread” use of ARs in urban areas (Cypher et al., 2014). Higher prevalence of 1101 commensal rodents which serve as vectors of ARs in urban areas may exacerbate this 1102 problem. A study in Canada demonstrated a higher proportion of rats in the diet of Barn 1103 Owls with territories containing more urban land use (Hindmarch and Elliott, 2014). 1104 Assuming that commensal rodents are an important vector of ARs, their higher relative 1105 proportion in the diets of urban owls may increase the incidence and severity of AR 1106 exposure. Boobooks are likely to be affected by this dynamic. In Canberra, Australia, 1107 boobook diets contained a higher percentage of mammal biomass in suburban areas 1108 (65.8%) than in woodland areas (26.0%) (Trost et al., 2008). Both the high prevalence of 1109 rodenticide use and the greater availability of potentially exposed commensal rodents likely 1110 contribute to the positive correlation between rodenticide exposure and developed habitat 1111 observed in boobooks. 1112 A negative correlation between AR exposure and the proportion of bushland area 1113 within simulated home ranges was expected because rodenticides are seldom used in native 1114 habitats, aside from the use of pindone to control rabbits. Only one other study has tested 1115 spatial patterns of AR exposure in wildlife primarily using bushland habitats. Unlike patterns 1116 observed in boobooks, high exposure rates were unexpectedly detected in fishers (Martes 1117 pennanti) throughout areas of forested habitat, probably as a result of rodenticide use 1118 associated with illegal marijuana production (Gabriel et al., 2012). Similarly a threatened 1119 Spotted Owl (Strix occidentalis) with illegal marijuana cultivation within its home range was 1120 documented to have been exposed to brodifacoum despite being in a remote natural area 1121 (Franklin et al., 2018). Conservation and law enforcement professionals should be aware of 1122 this potential source of environmental contamination when attempting to mitigate damage 1123 caused by illegal marijuana cultivation in remote areas in Australia. Future work examining 1124 the distance rodenticides travel into bushland ecosystems from adjacent sources will be 1125 useful in gaining a better understanding of the relationship between fragmentation and 1126 rodenticide use. This could potentially lead to establishing appropriate sizes for SGAR 1127 exclusion zones around bushland areas containing sensitive fauna and reduce edge effects 1128 relating to SGARs.

1129 The negative correlation between total AR exposure and the proportion of 1130 agricultural area within simulated home ranges was somewhat surprising, as rodenticides 1131 are known to be used in agricultural settings. AR exposure in wildlife has been attributed to 1132 agricultural application of ARs in the UK (Birks 1998; Hughes et al. 2013), Spain (Lemus et al., 1133 2011), France (Fourel et al., 2018), and Australia (Young and De Lai, 1997). Anecdotal 1134 accounts from farmers indicate that a variety of first and second generation products are 1135 used for asset protection around buildings and in grain storage areas in Wwestern Australia 1136 (Don Thompson personal communication). However, they are not licensed for use directly 1137 in crops or along crop perimeters. As a consequence, the total amount of bait deployed per 1138 unit area is likely to be substantially lower than in developed areas. However, in agricultural 1139 systems, total compliance with best practice application methods for SGARs may be rare 1140 and lack of compliance probably facilitates greater risk of secondary toxicity to native 1141 wildlife (Tosh et al., 2011). An anecdotal report of farmers in Western Australia requesting 1142 the FGAR pindone to control kangaroos (Twigg et al., 1999) – a use not allowed by the 1143 labelling – suggests that illegal use of ARs in agricultural contexts may be an issue in some 1144 areas. The widespread availability of SGARs to the public in Australia increases the risk that 1145 misuse could lead to localised impacts on non-target wildlife. 1146 The negative correlation between proximity to agricultural land and AR exposure 1147 may not be consistent throughout all Australian agricultural systems. In Queensland, 1148 declines in breeding owl abundance were attributed to broad-scale application of a 1149 brodifacoum-based rodenticide in canefields (Young & De Lai 1997) but this product was 1150 subsequently removed from the market (Twigg et al., 1999). At present, brodifacoum is 1151 only registered for use in and around buildings in Australia (McLeod & Saunders 2013) but 1152 can be freely purchased and applied without a license. While less toxic and persistent than 1153 brodifacoum, a coumatetralyl-based product is currently licensed for use in sugar cane, 1154 pineapple, and macadamia crops across Australia (Australian Pesticides and Veterinary 1155 Medicines Authority, 2017b). More concerningly, during rodent plagues the SGAR 1156 bromadiolone has been used to bait field perimeters in New South Wales (New South Wales 1157 Department of Primary Industries, 2011; New South Wales Government: Department of 1158 Primary Industries, 2017).

1159 Seasonal Differences 1160 The difference in AR exposure observed between boobook carcasses recovered in 1161 winter and those recovered in summer potentially reflects increased risk of exposure during 1162 winter when rodents make up a larger proportion of the diet. Boobooks are dietary 1163 generalists and one study indicates that boobook diet varies seasonally and includes higher 1164 proportions of vertebrates in winter than in autumn (Trost et al., 2008). This seasonal 1165 variation in diet may reduce the risk of accumulating lethal levels of ARs in boobooks 1166 relative to some other raptor species. Species preying predominantly on small mammals 1167 are likely to be at greater risk of exposure than species that prey predominantly on birds 1168 (Ruiz-Suárez et al., 2014). This hypothesis is supported by a lack of seasonal variation in AR 1169 exposure in Tawny Owls (Strix aluco) which feed consistently on bank voles (Myodes 1170 glareolus) and field mice (Apodemus spp.) (Walker et al., 2008). Similarly, in the United 1171 States, rodenticide exposure rates and concentrations did not vary significantly by season in 1172 Red-tailed Hawks (Buteo jamaicensis) (Stansley et al., 2014) which feed predominantly on 1173 mammals year-round. The only other study detecting seasonal variation in liver AR 1174 concentration found a significant difference in only one of five ARs tested (Christensen et 1175 al., 2012). This difference was attributed to an influx in autumn of migratory raptors from 1176 more sparsely populated regions with presumably less AR exposure risk (Christensen et al., 1177 2012). 1178 It is possible that consuming few rodents during a portion of the year allows 1179 boobooks to excrete sufficient levels of highly-persistent SGARs that total liver 1180 concentrations are less likely to accumulate to a lethal level. In this scenario, other raptor 1181 species which consistently consume rodents throughout the year – such as Masked Owls 1182 and Barking Owls – may be at elevated risk of lethal poisoning relative to boobooks. 1183 Alternately, seasonal variation in rodenticide exposure in boobooks could be correlated with 1184 seasonal differences in rodenticide use patterns. Information on rodenticide sales is not 1185 publicly available, but anecdotal accounts from some Perth residents indicate greater use of 1186 rodenticides in winter in response to greater perceived abundance of commensal rodents. 1187 Improved knowledge of rodenticide application patterns and seasonal patterns of 1188 rodenticide exposure in species with a more consistent mammal-based diet would be useful 1189 in addressing these questions.

1190 The high AR exposure rates observed in boobooks despite seasonal variation in the 1191 proportion of rodents in their diet highlights the need for additional study of exposure rates 1192 of other taxa which may potentially vector rodenticides. Documented exposure in raptors 1193 which prey primarily on birds indicates that non-rodent vectors may substantially contribute 1194 to AR exposure at higher trophic levels (Thomas et al., 2011). Invertebrates have been 1195 implicated in vectoring lethal levels of rodenticides to bird species including New Zealand 1196 Dotterels (Charadrius obscurus aquilonius) (Dowding et al., 2006) and nestling Stewart 1197 Island robins (Petroica australis rakiura) (Masuda et al., 2014) as well as an insectivorous 1198 mammal, the European hedgehog (Erinaceus europaeus) (Dowding et al., 2010). Reptiles 1199 could potentially also be effective vectors to higher trophic levels (Lohr and Davis, 2018). 1200 Further investigation of AR residues across more taxa is necessary to fully understand 1201 ecosystem-wide AR contamination and the vectors by which carnivorous species are 1202 exposed.

1203 Rodenticide in fledglings 1204 The detection of SGAR exposure in recent fledglings provides a possible indication as 1205 to why there was no significant difference in total AR exposure between hatch year 1206 boobooks and older adults. AR exposure prior to leaving the nest is particularly concerning 1207 from a conservation perspective. Suspected brodifacoum poisoning was previously 1208 reported as the likely cause of death of Norfolk Island Boobook chicks which were still in the 1209 nest (Debus, 2012) but there was no indication of physical examination or direct testing for 1210 AR exposure. Birds with growing feathers may be at additional risk of exsanguination 1211 (Newton et al., 1990). This may put chicks and recent fledglings at greater risk than adult 1212 birds which do not typically moult large proportions of their feathers simultaneously. 1213 Additional sub-lethal threats to chicks have also been reported. Stunted growth across 1214 several biometric measurements of nestling Barn Owls was observed in plots treated with 1215 anticoagulant rodenticides relative to control plots in Indonesia (Naim et al., 2010). While 1216 reduced prey availability due to rodent control likely had a negative influence on growth 1217 rates, nestlings in areas treated with the SGAR brodifacoum showed reduced growth when 1218 compared to areas where rodents were controlled with the FGAR warfarin or a biological 1219 rodent control agent (Naim et al., 2010), suggesting that AR exposure contributed to 1220 reduced nestling growth. Similarly, a dramatic reduction in breeding success occurred in a 1221 population of closely-related moreporks on Mokoia Island in New Zealand in the breeding

1222 season immediately following a broad-scale distribution of brodifacoum as part of an 1223 attempted mouse eradication (Stephenson et al. 1999). While Stephenson et al. (1999) 1224 concede that the reduction in breeding success may have been related to a drop in prey 1225 availability rather than a direct effect of rodenticide toxicity, depression of breeding success 1226 by anticoagulant rodenticides is plausible. Laboratory testing also detected modest 1227 reductions in weight gain and wing growth in juvenile Japanese Quail (Coturnix coturnix 1228 japonica) exposed to sub-lethal doses of brodifacoum or difenacoum (Butler, 2010). 1229 Perhaps the most conclusive evidence of negative impacts of sub-lethal AR exposure on 1230 growing birds is the correlation observed between concentrations of bromadiolone in blood 1231 and reduced body condition observed in nestling Common Kestrels (Falco tinnunculus) 1232 (Martínez-Padilla et al., 2016). 1233 Nest success may also be impacted in the early stages of nesting. Embryo toxicity 1234 has been observed in domestic chicken eggs injected with the anticoagulant rodenticide 1235 flocoumafen (Khalifa et al., 1992). It is also possible that exposure to anticoagulant 1236 rodenticides could impact egg viability via reductions in the integrity of eggshells. Exposure 1237 to therapeutic anticoagulants has resulted in bone density loss in humans by disruption of 1238 the vitamin K cycle and resultant suppression of calcification (Fiore et al. 1990; Resch et al. 1239 1991; Monreal et al. 1991) though similar effects on bone density have not been observed 1240 in birds (Knopper et al., 2007). Residues of bromadiolone and chlorophacinone were 1241 detected in yolk and albumin of addled Barn Owl eggs in areas of palm plantations treated 1242 with rodenticides but no changes to eggshell thickness or morphology were detected (Salim 1243 et al., 2015). However, changes to barn owl egg morphology, reduced eggshell mass and 1244 decreased eggshell thickness have been observed when eggs contained higher 1245 concentrations of brodifacoum (Naim et al., 2012). While teratogenic effects of 1246 anticoagulant rodenticides are not widely reported in birds, one study suggested this 1247 possibility when the authors detected a single barn owl nestling in a plot treated with 1248 brodifacoum which failed to grow primary feathers and would have been unable to fly 1249 (Naim et al., 2010). Haemorrhage of oviducts in association with rodenticide poisoning has 1250 been observed in female raptors carrying eggs (Murray, 2017), suggesting that ARs may 1251 pose a particular risk to nesting females. Future assessments of population-level impacts of 1252 anticoagulant rodenticide exposure need to consider not only adult mortality, but also 1253 impacts on fecundity and recruitment.

1254

1255 Conclusion

1256 My hypothesis that total AR exposure would vary between areas predominated by 1257 different types of anthropogenic landscape is to some degree supported by the finding of 1258 significant, though weak, relationships trending in opposite directions between total liver AR 1259 concentration and proportions of agriculture and developed land at the spatial scale of a 1260 boobook’s home range. Understanding this dynamic is key to assessing landscape-level risk 1261 of AR poisoning across carnivores and scavengers in Australia. It will also facilitate future 1262 attempts to model exposure risk in endangered and priority taxa which may be susceptible 1263 and will enable more specific risk assessment prior to proposed future developments. The 1264 high rates and magnitude of AR exposure raise serious concerns about AR exposure in other 1265 Australian species. Future work should evaluate the impact of ARs on other Australian 1266 wildlife, particularly species utilizing urban and peri-urban areas, species with large home 1267 ranges, and species regularly consuming commensal rodents. The detection in boobooks of 1268 ARs presumed to be used only by professionals is concerning. Ongoing review of the 1269 registration of SGARs by the APVMA should take this into consideration when evaluating the 1270 efficacy of restricting SGARs to licensed pesticide applicators in reducing poisoning in non- 1271 target wildlife. 1272

1273 Acknowledgements

1274 This project was supported financially by The Holsworth Wildlife Research 1275 Endowment via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird 1276 Research Award, the Edith Cowan University School of Science Postgraduate Student 1277 Support Award, the Eastern Metropolitan Regional Council’s Healthy Wildlife Healthy Lives 1278 program, the Society for the Preservation of Raptors, and Sian Mawson. 1279 I would like to extend my appreciation to the staff of National Measurement 1280 Institute’s (NMI), Analytical Services Port Melbourne branch, in particular Hao Nguyen and 1281 her veterinary drug residue measurement team for their support in providing high quality 1282 measurements using NATA accredited LC-MSMS measurement techniques. 1283 Cheryl Lohr provided valuable assistance in statistical analysis and Shaun Molloy 1284 graciously volunteered time to help develop necessary GIS layers. I would especially like to

1285 thank Kanyana Wildlife Rehabilitation, Native Animal Rescue, Native ARC, and Nature 1286 Conservation Margaret River Region and the many other individuals especially Simon 1287 Cherriman, Angela Febey, Amanda Payne, and Stuart Payne for contributing samples. This 1288 manuscript was improved by comments from Dr. Robert Davis, Dr. Allan Burbidge and two 1289 anonymous reviewers. The photograph in the graphical abstract was provided by Simon 1290 Cherriman. 1291

1292 Chapter 4 Widespread genetic connectivity in Australia’s most

1293 common owl, despite extensive habitat fragmentation

1294 Abstract 1295

1296 Lohr, M. T., P. B. S. Spencer, S. Krauss, J. Anthony, A. H. Burbidge, and R. A. Davis. 1297 Widespread genetic connectivity in Australia’s most common owl, despite extensive 1298 habitat fragmentation. The Condor: Ornithological Applications. (In Preparation).

1299

1300 Reductions in genetic diversity and genetic connectivity have been documented in 1301 some predatory bird species in response to anthropogenic habitat fragmentation. The 1302 Australian Boobook (Ninox boobook) is the most common and widely-distributed owl in 1303 Australia but declines in abundance have been observed across its range. To investigate 1304 whether genetic factors associated with habitat fragmentation have been associated with 1305 this reduction in abundance, we used polymorphic microsatellite loci to investigate patterns 1306 of genetic variation and its spatial structure in boobooks from a variety of fragmented and 1307 relatively undisturbed landscape types across Western Australia. The maximum distance 1308 between samples was 1391 km. Genetic analysis was informed by data on post-breeding 1309 dispersal of juvenile boobooks gathered from banding data resulting from this and other 1310 studies across Australia. We found weak spatial genetic structuring and no evidence of 1311 genetic erosion associated with inbreeding in heavily fragmented landscapes. Bayesian 1312 modelling and principal coordinates analysis suggested a single large panmictic population 1313 across all areas sampled. Within the heavily fragmented landscape of an extensive urban 1314 area, band re-sightings and recoveries substantiate the considerable capacity of juvenile 1315 boobooks to disperse across areas far greater than the distance between patches. We 1316 hypothesise that the genetic homogeneity observed is a consequence of long distance 1317 dispersal capacity of boobook offspring and their ability as habitat and dietary generalists to 1318 make use of highly altered habitats.

1319 Introduction

1320 Habitat Fragmentation, Connectivity, and Genetic Structure 1321 Habitat fragmentation can directly negatively impact the genetic diversity of a 1322 population of organisms by restricting gene flow between habitat patches and reducing 1323 effective population size (Aguilar et al., 2008). Reduction of effective population size can 1324 increase additional risks to small populations posed by demographic stochasticity, genetic 1325 drift, and inbreeding depression (Soulé and Simberloff, 1986). In fragmented environments, 1326 highly mobile species are less likely to experience these phenomena than species with 1327 limited dispersal capacity because of greater gene flow between fragments (Bohonak, 1328 1999). However, differences between the types and severity of threats found within distinct 1329 types of habitat matrix may drive differences in matrix permeability, irrespective of the 1330 biogeographic variables and mobility of the species concerned (Collinge, 1996). 1331 Simultaneous investigation of genetic connectivity in a single species across multiple matrix 1332 types has the potential to inform our understanding of the impacts of different matrices on 1333 permeability and metapopulation dynamics.

1334

1335 Genetic Responses of Predatory Birds to Fragmentation 1336 Predators are more frequently extirpated as a result of fragmentation than animals 1337 at lower trophic levels, as a result of their larger home range requirements and smaller 1338 population sizes (Didham et al. 1998; Gilbert et al. 1998; Duffy 2003). Predatory birds 1339 specifically have been observed to be at greater risk of extinction as a result of habitat 1340 fragmentation than other bird species (Leck 1979; Brash 1987; Carrete et al. 2009). This 1341 relative sensitivity to fragmentation makes predatory birds useful bio-indicators of 1342 ecosystem health in fragmented landscapes (Rodríguez-Estrella et al., 1998). A variety of 1343 negative impacts on predatory bird populations have been documented in association with 1344 use of highly fragmented landscapes, with some differences noted between urban and 1345 agricultural landscapes. In urban landscapes, documented negative impacts include 1346 increased mortality associated with electrocutions and collisions with vehicles and 1347 anthropogenic structures (Hager, 2009), reduced nest success due to higher rates of 1348 parasitic infection (Boal and Mannan, 1999), and higher rates of exposure to anticoagulant 1349 rodenticides (Lohr, 2018). Likewise, agricultural intensification has led to declines in

1350 carnivorous bird abundance as a consequence of loss of nesting sites, pesticide poisoning, 1351 and overgrazing of prey species habitat (Newton, 2004) as well as continental-scale decline 1352 across farmland bird species generally (Donald et al. 2001). While differences in landscape 1353 structure and the threatening processes associated with the urban and agricultural matrix 1354 clearly differ, it is unclear how or whether the different pressures exerted on predatory 1355 birds existing in these landscape types drive differences in matrix permeability and genetic 1356 connectivity.

1357 Relatively few studies have been conducted on the genetic impacts of habitat 1358 fragmentation on predatory birds, particularly across different types of anthropogenic 1359 matrix. In one instance, European Kestrels (Falco tinnunculus) were found to have greater 1360 relatedness in urban individuals as compared to rural individuals, despite similar allelic 1361 diversity in the two populations (Riegert et al. 2010) and genetic differentiation between 1362 urban and rural populations (Rutkowski et al. 2006). Within owls specifically, studies have 1363 not addressed the relative impacts of different matrix types but some have examined 1364 impacts of anthropogenic habitat fragmentation generally. Mediterranean Eagle Owls 1365 (Bubo bubo) have shown evidence of substantial population structure within a small 1366 geographic area of Spain as a consequence of anthropogenic habitat fragmentation (León- 1367 Ortega et al., 2014). Additionally, closely related individuals have been found in mated pairs 1368 of Powerful Owls in urban fringe areas but not in adjacent intact woodlands (Hogan and 1369 Cooke, 2010). Reductions in genetic diversity and connectivity in susceptible taxa like 1370 predatory birds may serve as an early indicator of ecosystem decay in fragmented 1371 landscapes. Further investigation of these factors has the potential to identify landscapes at 1372 risk of reductions in biodiversity at higher trophic levels as a consequence of extinction debt 1373 incurred via habitat fragmentation.

1374

1375 Declines in Australian Boobook Abundance 1376 The Australian Boobook (Ninox boobook) is a common and widespread owl species 1377 found across most of continental Australia but apparent range-wide declines have 1378 prompted calls to investigate potential drivers of reductions in abundance (BirdLife 1379 Australia, 2015). Consistent trends in historical accounts of boobook abundance support

1380 the hypothesis that they have specifically declined in abundance in the Perth metropolitan 1381 area in Western Australia. Alexander (1921) described boobooks as resident in Perth and 1382 referred to them as “the common owl of the district.” Serventy (1948) repeated Alexander’s 1383 assessment but with an added qualifier: “the common owl of the district, but not frequently 1384 heard in the immediate vicinity of Perth.” Several decades later, Storr & Johnstone (1988) 1385 described the boobook as a moderately common passage migrant in Perth with no breeding 1386 records but “possibly also an uncommon resident.” Most recently, Stranger (2003) directly 1387 suggested a decline in boobook abundance in urban areas of the Swan Coastal Plain, stating 1388 that they “formerly ranged broadly over the plain, but [are] now rarer in the suburbs.” 1389 While these statements are only qualitative, they paint a picture of a population in decline 1390 in conjunction with increased urbanization. A similar account from agricultural landscapes 1391 inland of Perth suggests a reduction in boobook abundance in the Shire of Northam 1392 coinciding with extensive clearing of bushland for agriculture in the 1930s: “Uncommon, 1393 widespread in small numbers but not heard as often as during 1930s” (Masters and 1394 Milhinch, 1974). A more quantitative study has also demonstrated a negative correlation 1395 between boobook abundance and intensity of urban development (Weaving et al., 2011). 1396 While secondary exposure to anticoagulant rodenticides likely explains some of the 1397 decrease in observations in urban and peri-urban environments since the 1980s (Lohr, 1398 2018), the drivers of this pattern in other areas of the country are not clear and exploration 1399 of potential genetic impacts of fragmentation is warranted. While all of these observations 1400 indicate declines related to conversion of natural landscapes to human land uses, it is 1401 unclear whether these responses are related to fragmentation or simply the loss of usable 1402 space.

1403

1404 Boobook Movement and Responses to Fragmentation 1405 At present, a variety of conflicting views on boobook movement and dispersal 1406 patterns exist. Most sources suggest they are year-round residents where they occur 1407 (Saunders & Ingram 1995; Higgins 1999; König et al. 1999), especially in cooler temperate 1408 areas (Olsen & Taylor 2001; Olsen et al. 2011). Within the southwest of Western Australia, 1409 Storr and Johnstone (1988) describe the boobook as a “passage migrant” on the Swan 1410 Coastal Plain, though other sources suggest that perceived migrations may merely reflect a

1411 decrease in detectability due to a reduction in calling during the non-breeding season (Olsen 1412 and Debus, 2013) or short-distance home range shifts by some females during the non- 1413 breeding season (Olsen and Taylor, 2001). McDonald and Pavey (2014) estimated 1414 movements of ≥ 32 km by boobooks in response to an arid zone rodent plague, potentially 1415 demonstrating longer-range temporary shifts in home range than suggested by Olsen and 1416 Taylor (2001). This estimate was based on the distance between their observations and the 1417 nearest assumed breeding habitat (Mcdonald and Pavey, 2014). These movements would 1418 have occurred across a landscape composed entirely of native bushland and may not be 1419 indicative of boobook movement patterns across fragmented landscapes. In summary, the 1420 existing evidence relating to boobook dispersal capacity across fragmented environments is 1421 limited and inconclusive.

1422 Across their range, boobooks are subject to predation by larger raptors, including 1423 Wedge-tailed Eagles (Aquila audax) (Cherriman, 2007) Grey Goshawks (Accipiter 1424 novaehollandiae) (Olsen et al., 1990), Brown Goshawks (Accipiter fasciatus) (Czechura et al. 1425 1987), and larger owls (Debus, 2009). Therefore, we hypothesised that they would be less 1426 likely to cross large sparsely-vegetated agricultural areas where they could be exposed to 1427 greater predation risk. That is, the matrix could be considered more hostile and less 1428 permeable in agricultural regions than in urban areas. Under these circumstances, 1429 fragmentation by agriculture in the wheatbelt could be functionally different to urban 1430 fragmentation in Perth with regard to dispersal and subsequent genetic impacts.

1431 Determination of genetic connectivity in boobooks via genetic analysis of individuals 1432 on a landscape scale will help settle long-standing speculation about the basic biology of this 1433 species and inform the management of an ecologically important and widespread avian 1434 carnivore. We aimed to determine whether potential differences in permeability of 1435 different types of anthropogenically-altered landscapes impacted genetic diversity and gene 1436 flow in a common but declining predatory bird by examining patterns of spatial genetic 1437 structure and corroborating these data with movement data derived from mark-recapture 1438 studies. We predicted that barriers to gene flow would occur in both urban and agricultural 1439 landscapes but would be more apparent in habitat fragmented by agricultural land use due 1440 to reduced dispersal capacity across a more hostile matrix.

1441 Methods

1442 Juvenile Dispersal 1443 To directly assess boobook dispersal capacity across fragmented habitats, we captured 1444 boobooks as nestlings or recent fledglings within their natal territory. Each young boobook 1445 was fitted with an individually-numbered stainless steel band issued by the Australian Bird 1446 and Bat Banding Scheme (ABBBS) to allow subsequent identification if re-sighted alive or 1447 recovered dead. A total of 17 boobooks from seven family groups were captured and 1448 banded. Of these, five individuals were re-sighted or found dead elsewhere. Location data 1449 submitted to the ABBBS by members of the public were then used to estimate dispersal 1450 distances. We also accessed data from the ABBBS from other banding studies elsewhere in 1451 Australia. We only included records of healthy birds captured in the wild to avoid potential 1452 bias from records of rehabilitated birds which may have behaved abnormally or been 1453 released away from the location where they were found. We found only eight additional 1454 qualifying instances of boobooks in Australia being banded as juveniles or nestlings and 1455 subsequently being resighted. One of these records was removed because the boobook 1456 was later recovered dead and still in the nest, leaving 12 available records, including those 1457 generated by our study.

1458 Genetic Sample Collection 1459 Western Australia is the largest state in Australia and covers an area of 1460 approximately 2,529,875 km² and makes up roughly the western third of the continent of 1461 Australia. We opportunistically collected blood and tissue samples from across the entirety 1462 of the state (Figure 4.1). We attempted to focus collection effort on three areas: the Perth 1463 metropolitan area, adjacent areas of continuous bushland in the Perth Hills, and agricultural 1464 areas in the agricultural Wheatbelt region in the vicinity the town of Kellerberrin 1465 approximately 200km east of Perth, in order to examine genetic structure across three 1466 distinct habitat types.

1467

1468 Figure 4.1 Sample locations of genotyped Australian Boobooks (Ninox boobook) in Western Australia. (“metro” = urban and 1469 suburban areas of Perth represented by squares, “rural” = forested area surrounding the Perth Metropolitan area 1470 represented by an “x” , “Southwest WA” = forested areas to the south of Perth represented by triangles, “Wheatbelt” = 1471 highly-fragmented agricultural landscapes represented by crosses, and “other” = Goldfields and Pilbara regions, 1472 represented by black circles, ‘other’ = Goldfields and Pilbara regions of Western Australia).

1473 We used several methods to collect genetic information. Live Australian Boobooks 1474 were captured using a noose pole (Olsen et al. 2011) at night in conjunction with audio lures 1475 while conducting occupancy surveys across landscapes dominated by urban, bushland, and 1476 agricultural habitats. Additional wild boobooks were captured opportunistically using a 1477 noose pole when roosting individuals and family groups were reported by volunteers during 1478 the day. Blood was also collected from live boobooks held by wildlife rehabilitators along 1479 with information about where the boobook was originally found. Blood was drawn from the 1480 right jugular vein of each captured boobook using an insulin syringe with a 25G needle 1481 designed for subcutaneous use. In larger birds where more than a single capillary tube of 1482 blood is required, it is preferable to take blood from the right jugular vein, as this reduces 1483 handling time and risk of hematoma relative to sampling from the brachial vein (Owen, 1484 2011). The blood was refrigerated and allowed to coagulate for at least 24 hours prior to

1485 being centrifuged at 13000 RPM for 10 minutes. The resulting serum was removed for 1486 disease testing and the remaining material was frozen at -20°C for later genetic analysis.

1487 Additional samples were taken from boobook carcasses and shed feathers solicited 1488 from private citizens through BirdLife WA and a network of volunteers. Feathers were 1489 stored in paper envelopes at -20°C. All carcasses were stored frozen at -20°C until 1490 dissection when samples of muscle tissue were removed and stored in 100% ethanol for 1491 later analysis.

1492 Boobooks (n=137) were placed into one of six predefined regions based on 1493 similarities in geography and landscape type. The category ‘Exurbs’ (n=28) included 1494 individuals collected in areas immediately surrounding but not within the Perth 1495 Metropolitan area. ‘Perth Hills’ specimens (n=8) originated in an area of continuous forest 1496 east of Perth. Birds placed in the ‘Perth Metro’ category (n=71) originated in urban and 1497 suburban areas of Perth. Some boobooks were obtained from the Goldfields and Pilbara 1498 regions of Western Australia and were placed together in the ‘Remote WA’ (n=4) category. 1499 Boobooks from wetter, cooler, forested climates to the south of Perth were placed in the 1500 ‘Southwest WA’ (n=17) category. The ‘Wheatbelt’ (n=9) category included all boobooks 1501 from highly-fragmented agricultural landscapes in the WA wheatbelt.

1502

1503 Genetic Analysis 1504 Microsatellites are commonly used in population genetic studies, particularly in bird 1505 species (Moura et al., 2017) for the purpose of individual fingerprinting, determining 1506 parentage, and exploring genetic variation and its spatial structure (Guichoux et al. 2011). 1507 Twenty microsatellite loci have been described for the Powerful Owl (Ninox strenua) and 19 1508 of these markers have been shown to be polymorphic in Australian Boobooks (Hogan et al. 1509 2009). Hogan et al. (2009) suggested these markers would be useful in genetic studies of all 1510 Ninox species tested. We used a subset of 15 loci microsatellites developed by Hogan et al. 1511 (2009) and after optimisation, nine (Nst02, Nst08, Nst11, Nst13, Nst14, Nst15, Nst16, Nst18, 1512 and Nst19; Table 4.1) were used to examine whether connectivity differed between the two 1513 types of anthropogenic landscapes.

1514 Table 4.1 The characteristics of the primers from 15 microsatellite loci amplified in Australian Boobooks (Ninox boobook) 1515 from Western Australia using primers adapted from (Hogan et al. 2007, 2009).

GenBank Primer (5`→3`) Locus Accession Reverse Core repetitive unit no. Forward & fluorescent dye

Nst01 EF512147 TTTTTCGCTGTTATTCCAAGG GGACCTGAAAATGCTGGATG [GT]10

Nst02 EF512148 PET-GCCTTCCTTTTCTGCAATGA CATCATGAAATCACGGTTCTC [CATC]12

Nst03 EF512149 GGGCAATAGCGAGCTACTCA TTTTTCCTACTAGTTCAAATCATGGA [CA]9CG[CA]11

Nst04 EF512150 TCTCCAGCTGAGGTTGTCCT AAATTCCCCTTCACCAATCC [GT]9

Nst05 EF512151 ATCCCACTCCAAATCACCAG GCCATTTTATATGCCGTAAACC [GT]13

Nst07 EF512153 TGCAGCTGCTTCTTTCTGTT GGAGGGACCTATGAGTGTGC [CATC]10

Nst08 EF512154 6-FAM-ATCAGGGGTTTAGGGTTGGT GCAGGAAAGACAGCAGGAAC [TG]17

Nst09 EF512155 ACATGGGAGGCAAAACACTC GCTTGCATCTGAAACCCAGT [CATC]23

Nst11 EF512157 VIC-TAAGCCTCACAGGAAGCACA TTGCTATTAAAGAATAACTGTGTGAGA [CTAT]10

Nst13 EF512159 PET-ACAATGCCAGAGCGGTATTT TTGAGGATGGCAAGGATTTC [CA]10GAGA CAGA[CA]9

Nst14 EF512160 TCTTCCTGAAGCCTGCAGAT TCCTCCCGTTTGTTCATTCT [CA]16

Nst15 EF512161 6-FAM-TCTGTGACTATCAGGCTGCTG CAGCACTGCAGGAAGATTGA [GT]8

Nst16 EF512162 PET-CCCAGAGATGTGCCTTCAGT GGCTGCCTGGTAGAAGATGA [CCAT]13

Nst18 EF512164 6-FAM-TTGCTTCAGTCATCCATCTGA TGTTTCCAAAAGCATAGAAAGAAA [AC]13

Nst19 EF512165 VIC-CAAGGCTGCTTTTCTTCCAA GCTCCAATCTATGAGCAGCA [AC]24 1516

1517 Australian Boobook DNA was extracted from either a 2mm2 piece of muscle tissue or 1518 50µl of blood using a salting out technique described by (Miller et al., 1988) and re- 1519 suspended in 100µl of amplification grade water.

1520 For amplification of microsatellite loci approximately 30 ng of genomic DNA was 1521 amplified by PCR with 5X polymerase buffer containing dNTPs (Fisher Biotec, Perth WA)

1522 either 1, 1.5 or 2 mM MgCl2 (Table 4.1), 0.2 µM F unlabelled primer and 0.4 µM of R primer 1523 and M13 labelled primer (Table 4.1), and 0.5 U Taq (Fisher Biotec) in a 10 µl reaction 1524 volume. PCR was performed on a Veriti thermocycler (Applied Biosystems). All samples 1525 were run on ABI 3500 Genetic Analyzer (Life Technologies) and scored using Genious V7.1 1526 (Biomatters, http://www.genious.com/). The following sets of loci were pooled together to 1527 run on the sequencer: (Nst02, Nst08, Nst11), (Nst13, Nst14, Nst15), (Nst16, Nst18, Nst19). 1528 Control samples were run in each PCR run to ensure compatibility between data used in the 1529 analysis.

1530 For loci Nst02 and Nst18, PCR conditions were an initial denaturation step at 94 °C 1531 for 5 minutes, followed by 30 cycles of denaturation at 94 °C for 30 seconds, annealing for 1532 45 seconds at 50 °C, and extension of 45 seconds at 72 °C with a final extension of 5 minutes 1533 at 72 °C. This was followed by another 8 cycles of denaturation at 94 °C for 30 seconds, 1534 annealing at 53 °C for 45 seconds and extension at 72 °C for 45 seconds. The last cycle was 1535 followed by final extension at 72 °C for 10 minutes. All other loci PCR conditions were an 1536 initial denaturation step at 94 °C for 5 minutes, followed by 4 touch down cycles of 1537 denaturation at 94 °C for 30 seconds, annealing for 45 seconds at 60-54 °C, and extension of 1538 45 seconds at 72 °C with a final extension of 5 minutes at 72 °C. This was followed by 1539 another 25 cycles of denaturation at 94 °C for 30 seconds, annealing at 54 °C for 45 seconds 1540 and extension at 72 °C for 45 seconds. The last cycle was followed by final extension at 72 1541 °C for 5 minutes. This was followed by another 8 cycles of denaturation at 94 °C for 30 1542 seconds, annealing at 53 °C for 45 seconds and extension at 72 °C for 45 seconds. The last 1543 cycle was followed by final extension at 72 °C for 10 minutes.

1544 Statistical Analysis 1545 Data for boobook owls were analysed at nine microsatellite loci described by Hogan 1546 et al. (2007) and Hogan et al. (2009). However, one locus was excluded from analysis 1547 (Nst14) due to a high frequency of genotyping failures, leaving eight microsatellite loci 1548 available for use in the results presented here. Highly related individuals, known offspring 1549 (sensu Wang 2018), and boobooks of unknown geographic origin were removed from 1550 analysis resulting in a sample size of 137 adult individuals (Appendix 4.1). We conducted 1551 Mantel tests and spatial autocorrelation using a subset of these individuals. Boobooks of 1552 known regional origin lacking precise location data were removed. Additionally, when more 1553 than one individual was sampled at a single location (usually in the case of mated pairs) one 1554 individual was randomly removed from analysis. This left 124 individuals available for the 1555 Mantel test. To examine spatial autocorrelation a subset of these individuals from the Perth 1556 Metro, Exurbs, and Perth Hills regions (n=98) were used because of the high density of 1557 sampling within these regions.

1558 To examine genetic relationships among groups of individuals and potential 1559 populations we conducted analysis of molecular variance (AMOVA) and principal 1560 coordinates analysis (PCoA) in GenAlEx6.502 (Peakall and Smouse, 2012, 2006). GenAlEx

1561 was also used to calculate descriptive statistics for each region including mean number of

1562 alleles (NA), effective number of alleles (NE), mean observed heterozygosity across all alleles

1563 (HO), mean unbiased expected heterozygosity across all alleles (uHE), and fixation index (F). 1564 We also used GenAlEx to examine trends in isolation by distance using a Mantel test using 1565 all individuals with known coordinates and another Mantel test including only boobooks

1566 from the Perth Metropolitan area. GenAlEx was also used to calculate pairwise FST, pairwise

1567 Jost’s DST, and Nm between regions. Spatial autocorrelation was tested in GenAlEx using 1568 even sample classes of n=200. We assessed genetic structuring using the program 1569 STRUCTURE 2.3.4 (Hubisz et al., 2009) using a burn-in of 100,000 steps and a MCMC of 1570 1,000,000 steps. We conducted 20 runs each assuming a different number of genetic 1571 clusters (K=1-6). We used CLUMPAK (Kopelman et al., 2015) to visually depict STRUCTURE 1572 outputs. STRUCTURE HARVESTER Web v0.6.94 (Earl and vonHoldt, 2012) was used to 1573 estimate the most probable number of genetic clusters using the Evanno et al. (2005) delta 1574 K method.

1575 Results

1576 Direct Measurement of Dispersal 1577 Across all 12 records, the average recorded distance between original capture 1578 location and subsequent observation in fledgling and nestling boobooks was approximately 1579 10.5km with a maximum recorded movement of 52 km (Table 4.2). In our study, juvenile 1580 boobooks were observed moving an average of 8km and up to 26 km from their capture 1581 site. All the captures and re-sightings of nestlings and fledglings from our study occurred 1582 within the Perth Metropolitan area across urban and suburban habitat.

1583 Table 4.2 Records of date a bird was tagged, its location, days and distances elapsed between capture and recovery of Australian Boobooks (Ninox boobook) banded as fledglings in Australia. 1584 Data from the Australian Capital Territory (ACT) and Queensland sourced from the Australian Bird and Bat Banding Scheme (http://www.environment.gov.au/science/bird-and-bat-banding). 1585 Western Australian data from re-sightings and recoveries of boobooks captured as part of this study.

Date Banded State/Territory Days elapsed Distance travelled (kms) Recovery Method between capture between capture and and recovery recovery 29-November-1993 ACT 62 0 Found on highway/road; but not certainly hit by car 30-June-1994 Queensland 154 52 Band number read in field (bird not trapped) 04-December-1994 ACT 106 8 Collided with a moving road vehicle 13-January-2000 ACT 1709 18 Found dead, cause unknown 20-January-2001 ACT 78 4 Collided with a moving road vehicle 03-January-2004 ACT 165 4 Found dead, cause unknown 14-February-2008 ACT 21 0 Found sick or injured 10-November-2015 Western Australia 46 0 Found sick or injured 08-December-2015 Western Australia 986 12 Found sick or injured 11-December-2015 Western Australia 167 0 Band number read in field (bird not trapped) 31-December-2015 Western Australia 16 2 Found dead, cause unknown 17-January-2016 Western Australia 125 26 Band number read in field (bird not trapped) 1586

1587

1588 Table 4.3 Analysis of Molecular Variance (AMOVA) results using six regional groups of Australian Boobooks (Ninox boobook) in Western Australia as populations.

Degrees of Sum of Mean Estimate of Source of variation freedom squares squares variance Variation (%) Among Populations 5 37.794 7.559 0.048 1% Within Populations 131 875.527 6.683 6.683 99% Total 136 913.321 6.731 100%

0.100 0.080 0.060 0.040

0.020

r r 0.000 -0.020 U -0.040 L -0.060 -0.080 3 7 9 11 13 15 17 19 21 23 25 27 29 31 33 35 38 42 46 53 65 Distance Class (End Point in kms) 1589

1590 Figure 4.2 A corellogram showing genetic correlation values (r) as a function of distance (kms) using eight microsatellite 1591 markers in a subset of Australian Boobooks (Ninox boobook) n=98 from the Perth metropolitan area, adjacent exurban 1592 areas and the Perth Hills. U and L are 95% confidence intervals around the null hypothesis of no spatial genetic structure. 1593 No significant genetic structure is shown at any distance class.

1594

1595

1596 Figure 4.3 Principal coordinate analysis results based on eight microsatellite loci in Australian Boobooks (Ninox boobook) in 1597 Western Australia. Clustering does not correspond to potential populations and is driven by two common alleles and their 1598 heterozygotes at the locus Nst15. Blue = 161/161, Green = 161/uncommon allele, Purple = 163/161, Orange = 1599 163/uncommon allele, Red = 163/163, Black = no result.

1600 Indirect Estimation of Dispersal 1601 All results suggested that there was no meaningful spatial genetic structuring in the 1602 population of boobooks we sampled. The Mantel tests did not detect a meaningful 1603 correlation between genetic and geographic distances in the entire group of boobooks

1604 sampled (Rxy=0.046, p=0.194) or within the metropolitan area (Rxy= 0.082, p=0.070). This 1605 result was corroborated by spatial autocorrelation analysis of boobooks from the Perth 1606 Metro, Exurbs, and Perth Hills regions which did not indicate significant genetic structure at 1607 any distance class (Figure 4.2). PCoA initially showed three distinct genetic clusters with no 1608 apparent correlation with hypothetical regions, and the first two axes explaining only 1609 15.91% of variance (Figure 4.3). Interrogation of the data set revealed that the three 1610 clusters were defined by homozygotes of two common alleles and their heterozygotes 1611 (Figure 4.3). When the locus Nst15 was removed from the analysis, no clusters were 1612 discernible (Figure 4.4) and the first two axes explained only 14.01% of the variance. The 1613 apparent clusters when the locus Nst15 was included appeared to be a consequence of a 1614 combination of low allelic diversity at the locus Nst15 and little genetic structure in the 1615 other seven loci. A lack of genetic structuring was also indicated by AMOVA which 1616 determined that 99% of the total molecular variance was partitioned within regions and 1617 only 1% among regions (Table 4.3). Fixation index values for all regions were within or below 1618 the range reported in populations of another small owl species which were not found to be

1619 impacted by a genetic bottleneck (Proudfoot et al., 2006) (Table 4.4). Pairwise Fst values 1620 between regions were low overall with the highest values between the Remote WA region 1621 and the other regions, consistent with largest geographic distance (Table 4.5). Similar 1622 patterns were evident in the estimated number of migrants per generation between regions 1623 (Table 4.5). However, even the highest values detected were still relatively low, particularly 1624 when taking into context the substantial geographic distance between the Remote WA 1625 collection locations and other regions and the large geographic area over which Remote WA

1626 specimens were obtained. Pairwise Jost’s DST values were also low between regions with 1627 the only significant value detected between the “Exurbs” and “Perth Metro” regions (Dst = 1628 0.027, P= 0.015) (Table 4.6). The statistical significance of this value is likely to be an 1629 artefact of the substantially larger samples sizes of these regions rather than indicative of a 1630 meaningful biological difference in the alleles present in the two regions. STRUCTURE 1631 results did not show any spatial genetic clustering (Figure 4.5). Low Delta K values also

1632 support a lack of spatial genetic structure (Figure 4.6). A single genetic cluster was 1633 supported by mean LnProb values obtained using CLUMPAK (Appendix 4.2) and STRUCTURE 1634 HARVESTER (Appendix 4.3).

Principal Coordinates (PCoA)

Exurbs

Perth Hills Perth Metro

Coord.2 Remote WA Southwest WA Wheatbelt

Coord. 1 1635

1636 Figure 4.4 Principal coordinate analysis results based on seven microsatellite loci (i.e. no Nst15 – see Fig 3) in Australian 1637 Boobooks in Western Australia. No clustering is apparent across or within six sampled regions (“Exurbs” = areas 1638 immediately surrounding but not within the Perth Metropolitan area, “Perth Hills” = an area of continuous forest east of 1639 Perth, “Perth Metro” = urban and suburban areas of Perth, ‘Remote WA’ = Goldfields and Pilbara regions of Western 1640 Australia, “Southwest WA” = forested areas to the south of Perth, “Wheatbelt” = highly-fragmented agricultural landscapes 1641 existing primarily between the “Remote” region and all other regions).

1642

1643 Figure 4.5 Visualization of Australian Boobooks (Ninox boobook) sampled from six regions in Western Australia (“Exurbs” = 1644 areas immediately surrounding but not within the Perth Metropolitan area, “Perth Hills” = an area of continuous forest 1645 east of Perth, “Perth Metro” = urban and suburban areas of Perth, ‘Remote WA’ = Goldfields and Pilbara regions of Western 1646 Australia, “Southwest WA” = forested areas to the south of Perth, “Wheatbelt” = highly-fragmented agricultural landscapes 1647 existing primarily between the “Remote” region and all other regions) using the STRUCTURE results from CLUMPAK 1648 comparing number of inferred genetic clusters (K) from 1-6. The data support a single genetic cluster. Each line represents 1649 an individual. The proportion of colours in each line represents the proportion of membership of each individual in each 1650 cluster.

1651

1652

1653 Figure 4.6 Plot of Evanno et al.’s (2005) delta K (ΔK) based on inferred genetic clusters (populations) ranging from 2 to 5 in 1654 Australian Boobooks (Ninox boobook) sampled from Western Australia.

1655

1656 Table 4.4 Genetic diversity parameters for Australian Boobooks (Ninox boobook) in six regions in Western Australia derived 1657 from eight microsatellite loci. Mean number of genotyped individuals (N), mean number of alleles per locus (NA), mean 1658 number of effective alleles (NE), mean observed heterozygosity (HO), mean unbiased expected heterozygosity (uHE).

Region N ± SE NA ± SE NE ± SE HO ± SE uHE ± SE Exurbs 27.3 ± 0.4 8.5 ± 1.5 4.48 ± 0.59 0.693 ± 0.049 0.747 ± 0.050 Perth Hills 7.9 ± 0.1 6.0 ± 0.8 4.53 ± 0.54 0.743 ± 0.077 0.793 ± 0.052 Perth Metro 69.6 ± 0.8 9.5 ± 1.5 4.51 ± 0.54 0.716 ± 0.047 0.755 ± 0.036 Remote WA 3.6 ± 0.3 4.4 ± 0.4 3.80 ± 0.41 0.875 ± 0.067 0.838 ± 0.031 Southwest WA 16.5 ± 0.2 7.8 ± 0.9 4.54 ± 0.62 0.688 ±0.069 0.767 ± 0.041 Wheatbelt 8.1 ± 0.2 5.6 ± 0.5 3.76 ± 0.38 0.769 ± 0.044 0.754 ± 0.037 1659

1660 Table 4.5 Pairwise Fst and estimated number of migrants per generation (NM) between all geographic regions of Australian 1661 Boobooks (Ninox boobook) sampled in Western Australia.

Region 1 Region 2 Fst Nm Exurbs Perth Hills 0.015 16.7 Exurbs Perth Metro 0.012 19.8 Perth Hills Perth Metro 0.017 14.1

Exurbs Remote WA 0.037 6.5 Perth Hills Remote WA 0.054 4.4 Perth Metro Remote WA 0.039 6.1 Exurbs Southwest WA 0.016 15.8 Perth Hills Southwest WA 0.027 9.1 Perth Metro Southwest WA 0.012 20.2 Remote WA Southwest WA 0.033 7.3 Exurbs Wheatbelt 0.030 8.1 Perth Hills Wheatbelt 0.036 6.6 Perth Metro Wheatbelt 0.019 13.2 Remote WA Wheatbelt 0.043 5.5 Southwest WA Wheatbelt 0.019 13.0 1662

1663 Table 4.6 Pairwise estimates of Jost's DST (below diagonal) and associated P values (above diagonal) for Australian 1664 Boobooks (Ninox boobook) sampled in five regions of Western Australia.

Exurbs Perth Hills Perth Metro Southwest WA Wheatbelt Exurbs 0.953 0.015 0.386 0.101 Perth Hills -0.049 0.679 0.562 0.485 Perth Metro 0.027 -0.016 0.235 0.363 Southwest WA 0.003 -0.011 0.011 0.752 Wheatbelt 0.041 -0.001 0.007 -0.028 1665

1666 Discussion

1667 Both direct (banding) and indirect (genetic analysis) estimates of dispersal indicated 1668 widespread connectivity across all sampled populations despite extensive historical clearing 1669 of bushland in urban and agricultural landscapes. All statistical tests performed indicate a 1670 single admixed population of boobooks across all areas sampled. This result is consistent 1671 with a previous study which showed very little phylogenetic distinction between putative

1672 boobook subspecies across continental Australia (Gwee et al., 2017). The slightly higher Fst 1673 values observed between boobooks in the “Remote WA” group and other groups are likely a 1674 consequence of the group’s small sample size and the large geographic area from which the 1675 samples were derived. Alternately, weak isolation by distance across a large geographic 1676 area could explain this trend.

1677 The weak spatial genetic structuring both across Western Australia and within and 1678 between fragmented habitats is likely caused by effective movement between remnant 1679 habitat patches by dispersing juveniles. The genetic connectivity observed between

1680 fragmented landscapes and adjacent intact landscapes suggests historical movement 1681 between all areas despite extensive clearing over a long period of time (Saunders, 1989) 1682 while the observed capacity in our banding studies, of juvenile boobooks to disperse across 1683 substantial distances within fragmented urban landscapes, demonstrates that this type of 1684 habitat alteration does not constitute a barrier to juvenile dispersal. This result is consistent 1685 with dispersal patterns observed in other owl species. In a telemetry study of Burrowing 1686 Owls (Athene cunicularia), fledglings dispersed an average of 14.9 km (range 0.2 km - 53.1 1687 km) from their natal nest (Rosier et al., 2006). Similar dispersal patterns were observed in 1688 Spotted Owls (Strix occidentalis) (LaHaye et al., 2001). Within Australia, congeneric 1689 Powerful Owls (Ninox strenua) have been observed dispersing up to 18 km from their natal 1690 nest across “urban fringe habitat” (Hogan and Cooke, 2010). Dispersal by juvenile boobooks 1691 of distances substantially greater than those between patches of bushland habitat provides 1692 a plausible explanation for the lack of genetic structuring observed in the boobooks tested. 1693 While only movements within regions were observed in this study, the long distance 1694 contemporary dispersal observed within the Perth Metro region suggests the capacity for 1695 substantial post-breeding dispersal between regions. This result is consistent with the 1696 genetic estimate of migrants per generation among regions, suggesting considerable 1697 historical dispersal of juvenile boobooks (Table 4.5).

1698 Additionally, in the course of the study, boobooks were frequently observed and 1699 captured in urban areas outside of remnant bushlands. In some instances boobooks were 1700 observed successfully fledging young in areas where their home range would be expected to 1701 encompass no bushland whatsoever and be composed entirely of moderately dense 1702 suburban housing and light commercial development. If highly anthropogenically-altered 1703 habitats are able to support successful breeding attempts, these habitats likely constitute 1704 usable space despite their high degree of alteration and would not constitute a barrier to 1705 dispersal. Detection of moreporks (Ninox novaeseelandiae) at 80% of bushland patches in 1706 an urban area in New Zealand (Morgan and Styche, 2012) and documented use of highly 1707 developed suburban habitat by a female boobook during the non-breeding season (Olsen 1708 and Taylor, 2001) supports the hypothesis that these highly-altered habitat types do not 1709 provide a barrier to dispersal in boobooks. It is unclear to what degree the majority 1710 components of agricultural landscapes are “usable habitat” for boobooks but, on one

1711 occasion, in the course of this study, a boobook was observed hunting along a road >1km 1712 from any bushland, tree line, or patch of native vegetation, suggesting that boobooks 1713 actively utilize resources in habitats which we initially hypothesized to function as a hostile 1714 matrix between patches of usable habitat. In two Australian passerine species, natural 1715 history traits associated with tolerance of the “hostile matrix” in a fragmented landscape 1716 were demonstrated to correlate with spatial patterns of genetic diversity (Shanahan et al. 1717 2011). Boobooks are generalist predators capable of utilizing a wide variety of habitat types 1718 and are clearly capable of juvenile dispersal across urban development. Their capacity to 1719 use a wide variety of habitat types including highly anthropogenically-altered landscapes 1720 likely facilitates connectivity across ostensibly “fragmented” habitat. The lack of resistance 1721 observed in fragmented landscapes in our study of booboks probably protects them from 1722 the negative genetic impacts of fragmentation. Recent modelling of Mexican Spotted Owl 1723 (Strix occidentalis lucida) gene flow across fragmented habitats suggests that landscape 1724 resistance was an important predictor of genetic distance between populations for species 1725 with high dispersal capacity in highly fragmented landscapes (Wan et al., 2018). Owl species 1726 with more specialised habitat and dietary requirements including Blakiston’s Fish Owls 1727 (Bubo blakistoni) (Omote et al., 2015), Spotted Owls (Strix occidentalis) (Haig et al., 2001), 1728 and the more closely related Powerful Owl (Ninox strenua) (Hogan and Cooke, 2010) have 1729 shown genetic bottlenecks and potentially dangerous levels of inbreeding in response to 1730 habitat fragmentation.

1731 The lack of evidence for inbreeding or isolation as a consequence of habitat 1732 fragmentation does not necessarily imply that populations of boobooks in landscapes 1733 fragmented by urban and agricultural developments are demographically healthy or self- 1734 sustaining. Weak spatial genetic structuring would likely also be observed in scenarios 1735 where fragmented habitats function as ecological sinks supported by healthy populations in 1736 adjacent intact habitats. This scenario is potentially even more likely in species with a high 1737 tolerance for altered habitats and substantial dispersal capacity. At least in urban areas, 1738 recent studies suggest that anthropogenic mortality from road strikes and secondary 1739 poisoning with anticoagulant rodenticides may pose significant threats to boobooks (Lohr, 1740 2018). Future work examining differences in life history parameters including adult and

1741 juvenile mortality across multiple habitat types would be useful in determining the relative 1742 utility of highly anthropogenically altered landscapes as boobook habitat.

1743 Genetic isolation and subsequent inbreeding could potentially become a problem for 1744 boobooks in urban and agricultural landscapes in the future despite their observed current 1745 dispersal capacity from banding studies if insufficient breeding hollows are retained at a 1746 landscape scale. Nest hollow availability is the key habitat requirement across the 1747 boobook’s range (Olsen and Taylor, 2001; Taylor and Canberra Ornitholgists Group, 1992) 1748 and urban fragments contain fewer hollow-bearing trees than intact forested areas (Harper 1749 et al. 2005). While nest hollow limitation does not currently appear to negatively impact 1750 boobooks in the Perth Metro area or WA wheatbelt (M. T. Lohr, unpublished data), 1751 continuing loss of nesting hollows through land clearing for additional development, 1752 inappropriate fire regimes, removal of nest trees for safety reasons, and urban infill could 1753 potentially reduce hollow availability in the future. In Powerful Owls (Ninox strenua), Hogan 1754 & Cooke (2010) detected instances of close inbreeding in two out of four pairs on the edge 1755 of urban areas near Melbourne despite a demonstrated capacity for dispersal up to 18km. 1756 Conversely, all three pairs nesting in continuous forested habitat were found to be 1757 unrelated (Hogan and Cooke, 2010). Hogan & Cooke (2010) speculated that this pattern 1758 could be explained by a lack of habitat for juveniles to disperse to, and subsequent 1759 clustering of related individuals, largely as a consequence of insufficient nest hollow 1760 availability. If patterns of boobook nest hollow availability ultimately approach those of 1761 Powerful Owls, this could lead to a reduction in genetic diversity and inbreeding depression 1762 over time in fragmented habitat types, even if boobooks are capable of dispersal between 1763 patches. However, if threatening processes and limiting factors in fragmented habitats are 1764 sufficiently addressed, both genetics and movement data suggest that boobooks should be 1765 capable of rapid recolonization and demographic recovery.

1766 Acknowledgments

1767 This project was supported financially by The Holsworth Wildlife Research 1768 Endowment via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird 1769 Research Award, and the Edith Cowan University School of Science Postgraduate Student 1770 Support Award. We thank Dr. Jamie Tedeschi for advice and technical assistance in 1771 laboratory work. We especially appreciate the contribution of boobook banding data by

1772 Jerry Olsen. Our research would not have been possible without contributions of samples 1773 and access to live birds provided by Kanyana Wildlife Rehabilitation, Native Animal Rescue, 1774 Native ARC, Nature Conservation Margaret River Region, Eagles Heritage Wildlife Centre, 1775 and many individual volunteers especially Simon Cherriman, Angela Febey, Amanda Payne, 1776 Stuart Payne, and Warren Goodwin.

1777 Appendix 4.A A complete listing of the samples used in the analysis of microsatellite DNA polymorphisms, 1778 including the identification number (Individual ID), sample source, collection dates, collection locations 1779 (decimal lat/long), sampling locations/regions and age at sampling of Australian Boobooks used in this study. 1780 HY=hatch year, SY=second year, AHY=after hatch year, ASY=after second year. Individual ID Sample Source Sample Day Sample Month Sample Year Latitude Longitude Region Age G0006 Red Cells 8 9 2015 -31.74144 115.978732 Exurbs HY G0007 Breast Muscle 7 3 2003 -31.88159 116.149899 Exurbs HY G0008 Red Cells 28 9 2015 -31.74849 115.778892 Perth Metro HY G0011 Red Cells 30 9 2015 -31.8015 115.804695 Perth Metro HY G0012 Red Cells 8 10 2015 -32.10015 115.790394 Perth Metro ASY G0013 Breast Muscle 20 7 2015 -32.0075 116.089187 Exurbs HY G0014 Pulled Feather 12 10 2015 -32.06889 115.800187 Perth Metro ASY G0016 Breast Muscle 21 10 2015 -31.71689 115.77686 Perth Metro SY G0018 Red Cells 23 10 2015 -31.67021 115.912591 Perth Hills HY G0020 Red Cells 8 11 2015 Southwest WA Unknown

G0021 Red Cells 8 11 2015 Southwest WA Unknown

G0022 Red Cells 8 11 2015 Southwest WA Unknown

G0025 Red Cells 8 11 2015 Southwest WA Unknown

G0027 Red Cells 10 11 2015 -31.81007 115.791618 Perth Metro AHY G0034 Red Cells 13 11 2015 -32.03368 116.31612 Perth Hills HY G0038 Red Cells 21 11 2015 -31.8015 115.804695 Perth Metro HY G0039 Red Cells 26 11 2015 -31.49202 117.73503 Wheatbelt ASY G0040 Red Cells 26 11 2015 -31.52239 117.72649 Wheatbelt ASY G0041 Red Cells 27 11 2015 -31.47847 117.68539 Wheatbelt Fledgling G0045 Red Cells 8 12 2015 -31.67703 115.71954 Perth Metro AHY G0047 Red Cells 9 12 2015 -31.96097 116.2743 Perth Hills ASY

G0049 Red Cells 11 12 2015 -31.7813 115.78518 Perth Metro SY G0054 Red Cells 15 12 2015 -31.94549 116.032607 Exurbs SY G0055 Red Cells 15 12 2015 -31.94549 116.032607 Exurbs ASY G0056 Red Cells 15 12 2015 -31.97075 115.81971 Perth Metro SY G0060 Leg Muscle 24 11 2015 -31.81009 116.000807 Exurbs SY G0061 Leg Muscle 19 12 2015 -34.09528 115.067182 Southwest WA Unknown G0062 Red Cells 31 12 2015 -31.79492 115.7495 Perth Metro SY G0065 Red Cells 4 1 2016 -31.95983 115.79725 Perth Metro SY G0066 Breast Muscle 31 12 2015 -31.99956 116.036 Exurbs Fledgling G0067 Red Cells 7 1 2016 -31.79374 117.66162 Wheatbelt SY G0068 Red Cells 10 1 2016 -32.21932 116.012926 Perth Metro HY G0070 Red Cells 10 1 2016 -31.89132 115.910063 Perth Metro ASY G0073 Red Cells 10 1 2016 Exurbs SY

G0075 Breast Muscle 15 12 2015 -34.10385 115.050051 Southwest WA HY G0076 Red Cells 17 1 2016 -31.9017 115.767664 Perth Metro Fledgling G0078 Breast Muscle 18 1 2016 -31.75827 116.002391 Exurbs Fledgling G0079 Red Cells 18 1 2016 -31.84031 115.80513 Perth Metro Fledgling G0080 Red Cells 18 1 2016 -31.96248 116.045246 Exurbs ASY G0081 Red Cells 18 1 2016 -31.96248 116.045246 Exurbs ASY G0082 Breast Muscle 20 1 2016 -31.19993 117.476303 Wheatbelt HY G0083 Breast Muscle 20 1 2016 -31.91668 115.857198 Perth Metro Fledgling G0084 Breast Muscle 27 1 2016 -31.90531 116.091333 Exurbs Fledgling G0085 Breast Muscle 30 1 2016 -31.79691 115.749024 Perth Metro Fledgling G0086 Breast Muscle 1 2 2016 -33.36447 115.683543 Southwest WA ASY G0087 Red Cells 2 2 2016 -31.86955 115.859803 Perth Metro Fledgling G0088 Red Cells 2 2 2016 -32.04082 115.9173 Perth Metro SY G0091 Red Cells 3 2 2016 -31.96032 115.82482 Perth Metro AHY G0093 Breast Muscle 4 2 2016 -31.928 115.834928 Perth Metro HY

G0094 Leg Muscle 5 2 2016 -31.75573 115.741913 Perth Metro Unknown G0095 Red Cells 8 2 2016 -31.76065 115.78703 Perth Metro ASY G0096 Red Cells 8 2 2016 -31.76065 115.78703 Perth Metro SY G0098 Breast Muscle 30 1 2016 -31.9173 116.058586 Exurbs SY G0099 Red Cells 10 2 2016 -32.01233 116.050934 Exurbs HY G0101 Red Cells 8 2 2016 -31.9109 115.8505 Perth Metro Fledgling G0102 Red Cells 11 2 2016 -32.04458 115.78187 Perth Metro HY G0103 Red Cells 10 2 2016 -31.54287 115.68851 Perth Hills ASY G0104 Breast Muscle 12 2 2016 -31.59496 115.701605 Perth Hills HY G0105 Red Cells 12 2 2016 -31.54907 115.6841 Perth Hills HY G0106 Breast Muscle 13 2 2016 -32.03654 116.104238 Perth Hills HY G0107 Breast Muscle 15 2 2016 -32.0093 116.064506 Exurbs HY G0108 Red Cells 18 2 2016 -31.92082 115.919792 Perth Metro HY G0109 Liver 27 2 2016 -32.33713 115.799745 Perth Metro HY G0110 Red Cells 1 3 2016 -31.93118 115.766318 Perth Metro HY G0111 Red Cells 2 3 2016 -32.01394 115.949982 Perth Metro Fledgling G0112 Red Cells 4 3 2016 -31.82104 116.141722 Exurbs Fledgling G0113 Breast Muscle 2 3 2016 -31.8795 115.95019 Perth Metro HY G0114 Breast Muscle 8 3 2016 -31.87583 115.800775 Perth Metro HY G0115 Red Cells 8 3 2016 -31.95804 116.052374 Exurbs HY G0116 Breast Muscle 21 2 2016 -32.03715 116.112922 Exurbs HY G0117 Liver 3 3 2016 -31.92284 115.759786 Perth Metro HY G0118 Breast Muscle 3 3 2016 -32.03494 115.883206 Perth Metro HY G0120 Breast Muscle 7 2016 -28.94821 114.780506 Wheatbelt HY

G0121 Leg Muscle 10 3 2016 -31.7626 115.809613 Perth Metro Unknown G0122 Breast Muscle 9 3 2016 -30.226 116.04 Wheatbelt HY G0123 Red Cells 15 3 2016 -31.79857 115.75175 Perth Metro AHY G0124 Other Deceased Tissue 15 3 2016 -31.79433 115.85868 Perth Metro Unknown

G0125 Breast Muscle 3 3 2016 -33.39701 115.648895 Southwest WA HY G0126 Breast Muscle 17 3 2016 -32.02177 115.798034 Perth Metro HY G0130 Breast Muscle 25 3 2016 -31.91323 115.939744 Perth Metro HY G0131 Breast Muscle 29 3 2016 -31.99207 115.904278 Perth Metro HY G0132 Red Cells 31 3 2016 -32.24146 116.00121 Exurbs HY G0133 Breast Muscle 30 3 2016 -32.04845 115.758353 Perth Metro HY G0134 Breast Muscle 30 3 2016 -31.9329 115.940048 Perth Metro SY G0135 Breast Muscle 31 3 2016 -31.77161 115.775272 Perth Metro HY G0136 Red Cells 2 4 2016 -31.97929 115.857025 Perth Metro HY G0137 Breast Muscle 4 4 2016 -31.94606 115.850553 Perth Metro HY G0138 Leg Muscle 7 2 2016 -32.12375 115.829529 Perth Metro Unknown G0140 Breast Muscle 25 4 2016 -31.89993 115.762612 Perth Metro HY G0141 Breast Muscle 25 4 2016 -31.93358 115.837546 Perth Metro HY G0142 Breast Muscle 11 3 2016 -33.99252 115.056734 Southwest WA HY G0143 Liver 9 5 2016 -31.80756 116.128456 Exurbs ASY G0144 Breast Muscle 16 5 2016 -32.04913 115.882706 Perth Metro HY G0145 Red Cells 23 5 2016 -31.92798 115.840554 Perth Metro HY G0146 Red Cells 23 5 2016 -31.9639 115.808737 Perth Metro HY G0147 Red Cells 23 5 2016 -31.86168 115.752841 Perth Metro Unknown G0148 Red Cells 25 5 2016 -32.03723 115.834908 Perth Metro HY G0150 Breast Muscle 30 5 2016 -32.00291 115.96533 Perth Metro HY G0151 Red Cells 2 6 2016 -32.05821 116.009846 Perth Metro SY G0154 Other Deceased Tissue 23 5 2016 -21.6675 116.2046 Remote WA Unknown G0155 Red Cells 17 6 2016 -31.99742 116.070711 Exurbs HY G0156 Red Cells 17 6 2016 -31.99742 116.070711 Exurbs ASY G0157 Red Cells 17 6 2016 -32.01666 115.936982 Perth Metro HY G0158 Breast Muscle 17 6 2016 -31.98453 116.054469 Perth Metro HY G0159 Breast Muscle 4 2015 -31.87568 116.216452 Exurbs HY

G0160 Breast Muscle 16 6 2016 -31.88085 115.978111 Perth Metro AHY G0161 Liver 24 5 2016 -32.36326 115.813895 Perth Metro HY G0162 Breast Muscle 6 7 2016 Perth Metro HY

G0163 Breast Muscle 13 6 2016 -27.696 114.67775 Remote WA HY G0164 Breast Muscle 15 8 2016 -31.96475 115.945326 Perth Metro HY G0165 Leg Muscle 8 2016 -31.8877 116.142588 Exurbs HY

G0166 Breast Muscle 20 7 2016 -31.82163 116.125182 Exurbs ASY G0167 Breast Muscle 13 9 2016 -31.98462 115.871278 Perth Metro HY G0168 Leg Muscle 9 9 2016 -31.65262 115.950282 Exurbs ASY G0170 Red Cells 6 10 2016 -31.88895 115.8792 Perth Metro ASY G0174 Leg Muscle 7 10 2016 -31.73355 115.825806 Perth Metro Unknown G0176 Breast Muscle 4 11 2016 -33.94818 115.417917 Southwest WA ASY G0177 Breast Muscle 9 2016 -26.22565 121.556821 Remote WA HY

G0178 Breast Muscle 18 7 2016 Perth Metro HY

G0179 Red Cells 16 11 2016 -31.88857 116.14066 Exurbs SY G0181 Breast Muscle 1 2 2017 -31.75482 115.810065 Perth Metro HY G0182 Leg Muscle 6 12 2016 -33.37924 115.684558 Southwest WA Unknown G0183 Breast Muscle 4 1 2017 -32.15944 115.818709 Perth Metro HY G0184 Breast Muscle 2 12 2016 -33.41889 115.70462 Southwest WA ASY G0185 Breast Muscle 12 2016 -31.96062 115.824735 Perth Metro HY

G0186 Breast Muscle 12 2016 Exurbs Fledgling

G0187 Breast Muscle 21 2 2017 -33.98397 115.088191 Southwest WA HY G0188 Breast Muscle 19 3 2017 -31.95911 116.095486 Perth Hills ASY G0189 Breast Muscle 24 3 2017 -31.97698 115.854673 Perth Metro HY G0191 Breast Muscle 13 5 2017 -33.95648 115.073144 Southwest WA HY G0192 Breast Muscle 5 7 2017 -31.98348 115.853238 Perth Metro HY G0193 Breast Muscle 2 7 2017 -33.68701 115.229887 Southwest WA Unknown G0194 Breast Muscle 18 4 2017 -32.18967 121.778519 Remote WA Unknown

G0195 Breast Muscle 5 7 2017 -31.03005 116.036572 Wheatbelt HY G0196 Breast Muscle 7 5 2017 -34.05445 116.169099 Southwest WA HY G0197 Breast Muscle 30 8 2017 -31.88769 116.595393 Wheatbelt Unknown G0198 Breast Muscle 26 10 2017 -34.1594 115.37132 Southwest WA Unknown

1781 Appendix 4.B CLUMPAK results showing median values of the natural 1782 log of the probability of the number of genetic clusters (K=1-6) in 1783 Australian Boobooks sampled in Western Australia.

1784 1785

1786 Appendix 4.C STRUCTURE HARVESTER output indicating the highest 1787 probability for K=1 in boobooks sampled in Western Australia. 1788

1789 Chapter 5 Artificial nest box supplementation does not affect

1790 Australian boobook (Ninox boobook) occupancy in fragmented

1791 habitats in south-western Australia

1792

1793 Lohr, M. T., S. Cherriman, A. H. Burbidge, and R. A. Davis. Artificial nest box supplementation 1794 does not affect Australian boobook (Ninox boobook) occupancy in fragmented habitats 1795 in south-western Australia. Wildlife Research. (In Review).

1796 Abstract 1797 Nest hollows are critical elements of usable habitat for many wildlife species 1798 worldwide, particularly in Australia. Loss of hollows due to anthropogenic processes and 1799 competition with introduced species over remaining hollows are key threats to hollow- 1800 nesting species in landscapes dominated by urban and agricultural development. 1801 Supplementation with artificial nest boxes has been suggested as a method to mitigate 1802 these threats but the efficacy of this technique has seldom been evaluated. The hollow- 1803 nesting Australian Boobook (Ninox boobook), a small owl, has experienced a nearly range- 1804 wide decline for reasons that are not well understood. We aimed to determine the utility of 1805 nest- box supplementation as a conservation action for boobooks and the influence of nest- 1806 box supplementation on potentially competing species across two different types of 1807 fragmented landscape. We monitored boobook occupancy in bushland fragments in urban 1808 and agricultural landscapes as well as in areas of continuous bushland before and after nest- 1809 box installation. Monitoring protocols involved nocturnal point counts and broadcast 1810 recordings of boobook calls and were based on methods used in previous owl surveys 1811 overlapping our study areas. We also used a pole- mounted video camera to record species 1812 using nest boxes during the boobook breeding season over the course of two years. Nest- 1813 box supplementation did not increase boobook occupancy at monitored sites over the 1814 period of this study, though one box was used successfully. Nest boxes were more 1815 frequently utilized by alien and overabundant native bird species. The ability of boobooks 1816 to use small hollows and possibly evict competing species probably insulates them from the 1817 impacts of hollow loss relative to other obligate hollow-nesting species.

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1818 Introduction

1819 Tree hollows are used for shelter and nesting by taxonomically diverse wildlife 1820 species worldwide and are often a critical component of habitat for those species (Martin 1821 and Eadie 1999; Isaac et al. 2014). As a result, a shortage of available hollows can limit 1822 abundance of hollow-nesting species (Newton, 1994). The loss of nest hollows is an issue of 1823 particular conservation concern in Australia with 302 species of vertebrate recorded as 1824 using hollows for shelter or nesting (Gibbons et al. 2002). Eleven per cent of all bird species 1825 in Australia are classified as obligate hollow nesters compared to 5% in Europe, 4% in North 1826 America, and 6% in Africa (Newton, 1994). The absence of primary hollow-excavating 1827 vertebrate fauna like woodpeckers (Picidae) in Australia leaves hollow formation primarily 1828 dependent on stochastic processes (Saunders et al. 1982) including fire, decay by fungal or 1829 insect attack, and mechanical damage from other trees, wind, or lightning (Fox et al. 2009). 1830 Termites play a major role in facilitating rot in heartwood and subsequent excavation of 1831 hollows throughout Australian woodlands (Gibbons et al., 2000). Hollow formation can take 1832 more than 150 years in some trees (Harper et al. 2005) but may occur more quickly in other 1833 tree species (Whitford, 2002). Tree hollows are currently being lost in Australia faster than 1834 they are being replaced (Lindenmayer et al. 1997). The confluence of these factors makes 1835 nest hollow availability a critical and possibly limiting factor in the habitat requirements of 1836 many Australian wildlife species.

1837 Fragmentation of woodlands by human land uses can increase the rate at which 1838 hollows and hollow-bearing trees are lost through a variety of mechanisms. In urban 1839 remnant woodlands, edge effects involving increased wind exposure may substantially 1840 impact the abundance of tree hollows and may impact the entirety of the fragment 1841 depending on its size (Harper et al. 2005). In a survey of tree hollow occurrence in urban 1842 remnant woodlands in Melbourne, Australia, Harper et al. (2005) found no hollows in 12 of 1843 44 survey sites and 64% of remnants contained fewer than six hollow-bearing trees per 1844 hectare which is “well below that contained in areas of un-logged non-urban forest”. Urban 1845 remnant forests in Sydney were also found to have fewer hollow-bearing trees than 1846 continuous forest (Davis et al. 2014). The removal of large trees for timber and firewood as 1847 part of past management practices has also substantially decreased the number of hollow- 1848 bearing trees in some urban remnants (Harper et al. 2005) and has directly impacted some

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1849 threatened bird species such as the Swift Parrot (Lathamus discolor) (Webb et al., 2018). 1850 The exclusion of fire from urban forest fragments and removal of large trees due to safety 1851 concerns, may also play a role in reducing hollow formation and persistence (Harper et al. 1852 2005). Conversely, the inappropriate use of fire has been noted as a key driver of hollow 1853 loss (Stojanovic et al., 2016) and in agricultural regions has combined with other stressors 1854 such as intentional bulldozing of nest trees, and lone trees in paddocks being blown over as 1855 a consequence of greater exposure to wind (Saunders et al., 2014).

1856 Consistent long-term decline in abundance of large nest hollows used by endangered 1857 Carnaby’s Black-Cockatoos (Calyptorhynchus latirostris) has also been observed in remnant 1858 bushlands in agricultural landscapes in Western Australia (Saunders et al., 2014). The 1859 relative paucity of available hollows in landscapes which have been intensively altered by 1860 humans may be a limiting factor for wildlife which would otherwise be capable of using 1861 remnant bushlands and could be a factor contributing to overall declines in biodiversity.

1862 Nest Competition and Predation 1863 Even where nest hollow abundance is high, competition from introduced and 1864 overabundant native species can reduce nest hollow availability for obligate hollow-nesting 1865 wildlife. In North America, range-wide decline in three bluebird species (Sialia spp.) has 1866 been partially attributed to competition for nest hollows from European Starlings (Sturnus 1867 vulgaris) and House Sparrows (Passer domesticus) (Newton, 1994). In Australia, the 1868 introduction of hollow-nesting Common Mynas (Acridotheres tristis) was found to be 1869 correlated with declines in three native hollow-nesting bird species in the Canberra area 1870 (Grarock et al. 2012). Introduced European honeybees have been recorded as excluding a 1871 wide variety of native marsupial species from nest boxes in Australia (Beyer and Goldingay, 1872 2006). Galahs (Eolophus roseicapilla) and Western Corellas (Cacatua pastinator) are native 1873 to Western Australia but are overabundant in some areas and are believed to negatively 1874 impact endangered Carnaby’s Black-Cockatoos through competition for scarce nesting 1875 hollows (Johnstone et al., 2015; Saunders and Doley, 2017). Predation by the introduced 1876 Sugar Glider (Petaurus breviceps) in Tasmania, Australia is the key cause of the decline of 1877 the endangered hollow-nesting Swift parrot (Stojanovic et al., 2014). Understanding 1878 interactions between native and introduced hollow nesting species will be of increasing

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1879 importance to conserving native biodiversity in areas where fragmentation simultaneously 1880 decreases hollow availability and facilitates growth of populations of introduced species.

1881 Impacts of Nest Boxes in Conservation 1882 As a response to hollow limitation, artificial nest hollow or “nest box” installation 1883 programs have been used for research aimed at understanding important life history traits 1884 of specific populations and have been used as an effective conservation measure to stabilize 1885 some declining populations (Lambrechts et al. 2012). These programs have been an 1886 important part of recovery efforts for hollow breeding birds worldwide. Routing of artificial 1887 hollows into living trees has been an integral part of successful efforts to increase 1888 abundance of the endangered Red-cockaded Woodpecker (Leuconotopicus borealis) in the 1889 southeastern United States (Walters, 1991). Widespread nest box provisioning efforts by 1890 private organizations have been widely attributed as a major factor in the recovery of three 1891 species of bluebirds in North America (Newton, 1994). Nest boxes have also been used to 1892 increase barn owl populations in Israel (Kan et al. 2013), Malaysia (Duckett and Karuppiah 1893 1990; Puan et al. 2012), and India (Parshad, 1999) as part of efforts to reduce crop damage 1894 by rodents. In Australia, construction of nest boxes is currently used successfully to mitigate 1895 losses of natural hollows for Carnaby’s Black-Cockatoos in the Western Australian 1896 agricultural zone (Johnstone et al. 2015) and Glossy Black-Cockatoos (Calyptorhynchus 1897 lathami) on Kangaroo Island (Mooney and Pedler, 2005) and has been used as a 1898 conservation tool in managing Critically Endangered Orange-bellied Parrots (Neophema 1899 chrysogaster) (Goldingay and Stevens, 2009) and Swift Parrots (Stojanovic et al., 2019). 1900 While most of these programs addressed lack of hollow availability, in some bird species, a 1901 variety of parameters impacting breeding success are higher in nest boxes than in natural 1902 hollows (Purcell et al. 1997).

1903 Nest boxes may not be a solution for all species, especially if nest hollow limitation is 1904 not the key cause of decline. For example, Loman (2006) found that nest hollow availability 1905 in small woodland patches was limiting for some obligate hollow-nesting passerine species 1906 but not others. In some instances, nest boxes may be preferred to natural nests and rapid 1907 adoption of nest boxes can give the appearance of nest limitation where there is none. For 1908 example, in one study, 83% of Tawny Owl pairs switched from natural nest sites to nest 1909 boxes within the year they were provided and 100% of pairs switched within four years but

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1910 breeding density did not appear to change as a result of nest box provisioning (Petty, 1992). 1911 Purple Martins (Progne subis) in North America provide an even more extreme example of 1912 this dynamic. The eastern population of Purple Martins has used nesting structures 1913 provided by humans since prior to European colonization (Speck, 1941) and is now almost 1914 completely dependent on artificial nesting hollows constructed by humans (Morton et al. 1915 1990). While human-provisioned nest hollows clearly benefit this species, the potential 1916 risks of a population’s near-complete dependence on nest sites provided by humans are 1917 evident. In instances where nest boxes are preferred to natural hollows but are associated 1918 with lower nesting success they may even function as ecological traps (Klein et al. 2007, 1919 Heinshohn et al., 2015). Perhaps most fundamentally, nest box supplementation will not 1920 result in increases in abundance of the target species unless other resource requirements 1921 are already met (Durant et al. 2009). These factors should be considered before 1922 implementing or encouraging large-scale nest box programs and when evaluating the 1923 results of these programs.

1924 Knowledge Gaps 1925 Despite a large body of research on nest box impacts on native mammals and use of 1926 nest boxes in conservation efforts for cockatoos, few studies have focused on use of nest 1927 boxes by predatory birds in Australia. In a review of literature regarding nest box use by 1928 Australian bird species, only one of 17 species listed as having been studied was a predatory 1929 bird (Goldingay and Stevens, 2009). This study was conducted on a small hybrid population 1930 of boobooks on Norfolk Island and was an overview of conservation efforts rather than an 1931 empirical study of nest box impacts (Olsen, 1996). Another major knowledge gap relating to 1932 nest box impacts involves their use in developed areas. Less than 5% of Australian studies 1933 on the use of natural and artificial hollows have been conducted in urban landscapes 1934 (Durant, 2006).

1935 Despite the lack of studies relating to use of nest boxes by urban birds generally and 1936 Australian raptors specifically, artificial nest hollows have already been promoted as a 1937 conservation measure for urban raptors. Provision of nest boxes was suggested to improve 1938 Powerful Owl habitat in urban environments where scarcity of suitable nest hollows may be 1939 limiting abundance (Isaac et al. 2008). In one instance, subsequent localized nest box 1940 placement resulted in successful breeding of a nesting pair (McNabb and Keating, 2008;

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1941 McNabb and Greenwood, 2011). While this particular effort was well justified and the 1942 result of this action is encouraging, the unregulated implementation of untested 1943 conservation actions intended to benefit sensitive species is concerning. Nest boxes 1944 intended for use by a wide variety of wildlife species are already commercially available 1945 from local businesses and instructions and plans are readily available online and are actively 1946 promoted for owls by the WA government Department of Biodiversity, Conservation and 1947 Attractions (Hussey, 1997). Both options are promoted as broadly beneficial to native 1948 wildlife despite a lack of rigorous testing for most species. Klein et al. (2007) suggested that 1949 correlation between increased breeding abundance and nest box provisioning should be 1950 proven prior to use of nest boxes as a conservation strategy. The widespread promotion 1951 and use of nest boxes necessitates studies addressing impacts of nest boxes on bird 1952 populations broadly and particularly on predatory birds and birds using urban areas.

1953 We studied the small owl, the Australian boobook in south-western Australia, as a 1954 model to examine whether nest box provisioning can increase occupancy by this species in 1955 human-altered landscapes. Australian boobook’s are an ideal study species as they are 1956 widespread but a 2015 report on population trends in Australian birds identified a serious 1957 decline in Australian boobook numbers from 1999-2013 and recommended that “further 1958 investigation is needed to understand the factors that are driving this consistent decline 1959 across regions” (BirdLife Australia, 2015). Nest hollow availability is believed to be the key 1960 habitat requirement across the boobook’s range (Olsen and Taylor, 2001; Taylor and 1961 Canberra Ornitholgists Group, 1992) and loss of tree hollows has been cited as one of the 1962 reasons for its decline in some areas (Debus, 2009). In the single published study involving 1963 nest box use by boobooks, lack of nesting hollows was implicated as one of the major 1964 factors contributing to the near extinction of Norfolk Island boobooks (Ninox 1965 novaseelandiae undulata) and nest boxes were a key tool used in its recovery program 1966 (Olsen, 1996). Boobook occurrence has been observed to correlate negatively with 1967 increased density of sealed roads and positively with forest cover, and nest hollow 1968 availability has been hypothesized as the factor driving differences in boobook abundance 1969 between urban and forested landscapes in and around Melbourne, Australia (Weaving et 1970 al., 2011). Urban fragments generally contain fewer hollow bearing trees than intact 1971 forested areas (Harper et al. 2005). Likewise, in the agricultural “wheatbelt” of Western

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1972 Australia, loss of nest hollows is said to be one of the most important challenges facing 1973 wildlife conservation (Johnstone et al. 2015). Examination of patterns of nest box use by 1974 boobooks and its relationship with site occupancy across these two habitat types is 1975 necessary to understand their potential utility in the conservation of this species.

1976 Specifically we aimed to investigate whether Australian boobook occupancy in 1977 fragmented landscape types (agricultural and urban) was altered by providing nest boxes. 1978 Our hypothesis was that nest hollows would be limiting in agricultural and urban landscapes 1979 and that nest boxes would be quickly taken up by Australian boobooks.

1980

1981 Methods

1982 Study Sites 1983 To determine the impacts of nest box installation on site occupancy, surveys were 1984 conducted in 2015 at >30 sites each in each of three categories of land use: urban remnant 1985 bushlands, agricultural remnant bushlands, and areas of continuous bushland . Sites were 1986 located along the same approximate latitude in an area of south western Western Australia 1987 with a Mediterranean climate (Figure 5.1). Urban sites (n=35) found across the Perth 1988 metropolitan area were composed of bushland reserves managed by city governments or 1989 the Botanic Parks and Gardens Authority. Most sites were open woodlands dominated by 1990 Banksia sp., Eucalyptus gomphocephala, or E. rudis. Agricultural sites (n= 33) included both 1991 privately-owned bushlands and sites managed by the Western Australian Department of 1992 Biodiversity, Conservation and Attractions. All were within approximately 60km of the town 1993 of Kellerberrin, Western Australia. Dominant vegetation across these sites included Acacia 1994 acuminata, Eucalyptus capillosa, E. loxophleba, and E. salmonophloia. Continuous bushland 1995 sites (n=34) were located between the Perth Metropolitan area and areas of extensive 1996 agricultural development. They were bounded by the Great Eastern and Great Southern 1997 Highways to the North and the Brookton Highway to the South. Dominant vegetation in 1998 these sites was primarily Eucalyptus wandoo, E. marginata, and Corymbia calophylla. Intact 1999 bushland sites were included in surveys as a baseline against which to compare the efficacy 2000 of nest box supplementation as a management action intended to increase site occupancy.

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2001

2002 Figure 5.1 Locations of survey sites in in southwestern Western Australia: urban landscapes in the Perth Metropolitan Area, 2003 continuous bushland in the Perth Hills, and agricultural landscapes within a 60km radius of Kellerberrin, Western Australia.

2004 Surveys 2005 In urban and agricultural bushlands, surveys were conducted 100m from a road or 2006 near the middle of the reserve in smaller reserves to reduce the impact of traffic noise on 2007 surveys. In continuous bushland areas, survey points were located approximately 5km apart 2008 to ensure independence. Baseline occupancy surveys were conducted in 2015 from 2009 September to December during the breeding season when boobooks call most frequently 2010 and are most easily detected (Olsen, 2011b). To maintain consistent detectability of 2011 boobooks, surveys were only conducted in the absence of rain and when estimations of 2012 wind speed were below a score of 3 on the Beaufort scale. Surveys consisted of passively 2013 listening for boobook vocalizations from a fixed point for 15 minutes followed by five 2014 minutes of intermittent broadcast of recorded boobook vocalizations in accordance with 2015 methodology used by Liddelow et al. (2002) to survey nocturnal birds in south-western WA. 2016 Immediately following the survey, the area was scanned using a 1000 lumen LED headlamp 2017 to detect any boobooks that had been attracted by the calls but had not vocalised. All sites

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2018 were classified as “occupied” or “not occupied.” A subsequent round of surveys was 2019 conducted at all sites in September-December of 2016 after the installation of nest boxes to 2020 determine occupancy using the same methodology used during the previous breeding 2021 season.

2022 Nest box construction and placement 2023 Nest boxes were constructed using recycled, 18mm form-ply, a waterproof and long- 2024 lasting material used mainly for concrete construction work. Each box consisted of a 2025 wooden cube measuring 300mm long and 300mm wide, and a depth ranging from 450mm 2026 at the front to 500mm at the rear, creating a forward-sloping roof. The dimensions of the 2027 box were chosen to reflect dimensions of active boobook nests observed by the authors and 2028 to deter Galahs, which prefer deeper boxes and are aggressive competitors for nest hollows. 2029 A hollow log-round of diameter 120-200mm was attached to the front of each box using 2030 screws fastened from the inside. This served to create a ‘verandah’ designed to protect the 2031 internal nest-chamber from weather, and also to prevent non-target species with 'heavy- 2032 chewing' behaviour (e.g. Galahs) from enlarging the box entrance hole and potentially 2033 destroying it. A wooden lid with a c. 50mm overhang was attached with a hinge fitted to the 2034 rear, and the sides were reinforced with aluminium flashing, again to prevent chewing 2035 species from destroying the lid. Boxes were assembled in such a way to leave ’air slots’ 2036 ~15mm wide beneath the lid on both sides, designed to facilitate air-flow and subsequent 2037 internal temperature fluctuation to deter feral honey bees, which have specific hive 2038 temperature requirements of 32-35˚C, from taking up residence. Two coats of pale-green, 2039 water-based exterior paint were applied to all external surfaces, to protect the sawn 2040 wooden edges from the elements and thus defer deterioration, and to help the boxes blend 2041 in with the natural environment. A layer of coarse woodchips c. 150mm deep was added to 2042 the inside of each box to create an internal nest chamber consisting of well-drained 2043 substrate that allows hollow-nesting species to scrape a shallow bowl in which eggs are 2044 deposited. Wood chips consisted of c. 20mm diameter pieces and were collected near 2045 installation sites.

2046 Nest boxes were installed in trees with multi-strand, galvanised wire (‘clothesline 2047 wire’) c. 4mm thick, threaded through plastic/rubber pipe (‘hosepipe’) to protect the tree's 2048 bark from wire damage. The installation process was carried out using the following steps:

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2049 1) a small loop was created at one end of the wire, and the tail end threaded into one of two 2050 c. 8mm holes pre-drilled on the rear surface of the box, at each of its top corners; 2) the tail 2051 end was then threaded out through the second hole and the wire pulled through until it was 2052 tight; 3) after being threaded through a length of hosepipe, the main length of wire was 2053 looped horizontally around a solid, vertical section of trunk, being passed above an oblique 2054 or horizontal limb used to ‘hang’ the box and prevent it sliding down (Figure 5.2); the tail 2055 end was then threaded through the small loop at the back of the box and twitched into 2056 place for secure attachment. Sufficient length of wire was used so each box was ’strung’ 2057 firmly but not hung in such a way that left wire tightly constricting on the trunk. This 2058 method is similar to the ‘habisure system’ described in Franks and Franks (2006), and it 2059 ensures secondary (horizontal) growth of the tree’s trunk (i.e. limb thickening) can take 2060 place naturally. Permanent attachment methods involving fixings such as coach bolts or 2061 screws were avoided to 1) minimise injury to the tree’s vascular cambium that may lead to 2062 unnecessary infection or damage, and 2) ensure boxes were not ‘pushed off' as the tree 2063 trunk expands during secondary growth, resulting in the potential collapse of an occupied 2064 nest-site and/or a safety risk to passers-by.

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2065

2066 Figure 5.2 Attachment system used to hang nest boxes used in this study.

2067 Nest boxes were placed in fifteen sites in both urban and agricultural remnant 2068 bushlands which did not have boobook detections in the previous round of surveys. Nest 2069 boxes were installed in February 2016 shortly after the termination of the breeding season 2070 to allow adequate time for detection by boobooks prior to the following breeding season. 2071 All nest boxes were placed in the nearest suitable tree to the survey point in all 30 2072 experimental sites. All nest boxes were hung at a height below 11m to facilitate observation 2073 of their contents and greater than 4m because most published records indicate minimum 2074 nest heights above 3m for boobooks (Higgins, 1999) (Figure 5.3).

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2075

2076 Figure 5.3 A nest box installed in one of the remnant bushlands in an agricultural landscape in Western Australia.

2077 Nest Box Monitoring 2078 We examined the contents of all nest boxes for evidence of use by boobooks or 2079 potentially competing species. Nest box contents were viewed using a video camera 2080 (MiGear ExtremeX Sports Action Camera) mounted on an 8m telescoping fiberglass pole to 2081 record video footage of the inside of each nest box. All videos were viewed at the nest site

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2082 to ensure that adequate footage of the nest boxes’ interior was obtained to allow 2083 identification of contents. Videos were retained for later review. Nest boxes were checked 2084 on three occasions during the breeding season in 2016 (July 24-26, October 7-9, and 2085 November 18-25) and once during the 2017 breeding season (September 27-29). In the 2086 2017 surveys, a single nest box at one of the urban sites was unavailable to be checked as it 2087 had been destroyed by a bushfire.

2088 Statistical Analysis 2089 We compared differences in territory occupancy in 2015 across all three habitat 2090 types prior to treatment using a Chi-square test with a post hoc pairwise test of 2091 independence for nominal data. We used McNemar's Chi-squared tests with continuity 2092 correction to examine differences in occupancy between years at treated and untreated 2093 sites. All tests were performed using RStudio 1.1.383 (RStudio, Inc., Boston, MA, USA).

2094 Results

2095 Prior to nest box treatment, boobooks were more commonly detected in continuous 2096 bushland sites (85.3%, n = 34) than in remnant bushlands in urban (30.8% n = 39) and 2097 agricultural (21.2%, n = 33) landscapes. Occupancy rates were significantly greater at 2098 continuous bushland sites than urban sites (p<0.001) or wheatbelt sites (p<0.001) but did 2099 not differ between urban and wheatbelt sites (p=0.517). No significant differences in 2100 occupancy were detected between years in any of the treated or untreated groups across all 2101 three habitat types (Table 5.1). However, non-significant increases in occupancy occurred in 2102 sites provided with nest boxes in both fragmented habitats while non-significant declines 2103 occurred in control sites in both urban and agricultural habitats (Table 5.1).

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2104

2105 Table 5.1 Annual change in occupancy of Australian Boobooks at continuous bushland sites and sites with and without supplemental nest boxes in remnant woodland in urban and agricultural 2106 landscapes in Western Australia.

No. Occupied % Occupied No. Occupied % Occupied McNemar's Total sites 2015 2015 2016 2016 chi-squared df p value Urban with box 15 0 0.0 1 6.7 0 1 1 Urban without box 24 12 50.0 7 29.2 1.7778 1 0.1824 Wheatbelt with box 15 0 0.0 3 20.0 1.3333 1 0.2482 Wheatbelt without box 18 7 38.9 6 33.3 0 1 1 Continuous bushland 34 29 85.3 29 85.3 0 1 1 2107

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2108 Table 5.2 Number of nest boxes used by bird species in urban and agricultural remnant woodlands across two years in 2109 Western Australia.

2016 2017 Urban Wheatbelt Urban Wheatbelt Australian Boobook (Ninox boobook) 1 Australian Wood Duck (Chenonetta jubata) 1 1 2 3 Laughing Kookaburra (Dacelo novaeguineae) 3 2 Australian Ringneck (Barnardius zonarius) 1 Butler’s Corella (Cacatua pastinator butleri) 1 Galah (Eolophus roseicapilla) 1 2110

2111 Nest boxes were used by a total of six species (Table 5.2). The most commonly 2112 detected species utilizing nest boxes was the Australian Wood Duck (Chenonetta jubata). 2113 The only exotic species observed nesting in the nest boxes was the Laughing Kookaburra 2114 (Dacelo novaeguineae; introduced to Western Australian in the early 1900s). All nest boxes 2115 used by this species were in urban bushlands. Boobooks used one nest box, located in one 2116 of the urban bushlands, during the 2016 breeding season. Three large and healthy owlets 2117 were observed in this nest box. We assumed this nest box to have been successful because 2118 the age of the nestlings calculated using the equation given by Olsen et al. (2015) was 2119 greater than the average age at which boobooks fledge.

2120 Discussion

2121 Surveys 2122 The detection of boobooks in 85.3% of continuous bushland areas in both years is 2123 higher than in previous surveys conducted in similar forested areas of Western Australia 2124 during spring (Liddelow et al., 2002), in which boobooks were detected at only 61.5% of 2125 sites. Our use of boobook calls, which were not used in previous studies, likely improved 2126 our ability to detect boobooks that were present. It is also possible that broadcasting the 2127 calls of Barking Owls and Masked Owls in previous studies may have supressed boobook 2128 calling, as these species are larger and may compete with or prey on boobooks. Boobooks 2129 will sometimes stop calling in response to broadcast calls of Powerful Owls or Masked Owls 2130 (Debus, 2009). A study of boobooks in suburban and forested habitats around Melbourne,

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2131 Victoria found higher boobook occupancy in forested sites (94%) than our study (85.3%) but 2132 substantially lower occupancy in suburban sites (13%) than was observed in our study prior 2133 to nest box supplementation (30.8%) (Weaving et al., 2011). The higher occupancy 2134 recorded by Weaving et al. (2011) in forested areas may simply be an artefact of greater 2135 sampling effort and the use of transects rather than point counts. However, the difference 2136 in occupancy rates in urban areas runs counter to what would be expected given the 2137 differences in methodologies, suggesting an actual difference. Some of the urban areas of 2138 Perth where our study was conducted have been developed more recently than the sites in 2139 Melbourne. This may mean that extinction debt generated as a result of fragmentation at 2140 urban sites in Perth has not been fully paid, which could explain the disparity in occupancy 2141 between the two studies. This hypothesis is consistent with the observation that despite 2142 nest box supplementation at some sites, boobook occupancy across all urban sites dropped 2143 from 30.8% to 20.5% between 2015 and 2016. If this hypothesis is correct and our 2144 occupancy estimates are low relative to those of Weaving et al. (2011) due to our fewer 2145 surveys and different methodology, further reductions in boobook abundance can be 2146 expected in the Perth Metropolitan area.

2147 While the decline observed in boobook occupancy at urban sites is difficult to 2148 substantiate due to low sample size and an insufficient number and duration of surveys at 2149 the same sites, it is consistent with national trends indicating a continental scale decline in 2150 boobook abundance (BirdLife Australia, 2015) and reductions in boobook occupancy in 2151 urban bushlands in Canberra (Olsen and Trost, 2015). Conversely, occupancy was roughly 2152 stable bushland fragments in agricultural landscapes and continuous bushland. Recent 2153 research on boobooks in Western Australia documented pervasive and sometimes lethal 2154 exposure to anticoagulant rodenticides associated with proximity to developed habitat, but 2155 not agricultural or bushland habitat (Lohr, 2018). Secondary anticoagulant poisoning is a 2156 plausible mechanism explaining the observed differences in population trajectories across 2157 these three landscape types.

2158 Nest Box Use 2159 The use of nest boxes primarily by introduced species and overabundant native 2160 species must be considered when evaluating the use of nest boxes as a tool in conservation. 2161 Laughing Kookaburras are not native to Western Australia and anecdotal accounts of

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2162 negative impacts on breeding native passerines (Serventy, 1980) suggest that facilitating 2163 breeding of Laughing Kookaburras should not be encouraged. While Laughing Kookaburras 2164 have not been documented directly competing with boobooks for nest hollows, they have 2165 been observed directly killing roosting boobooks during the day on several occasions 2166 (Higgins, 1999).

2167 Aside from boobooks, all native species documented using nest boxes in this study 2168 are subject to control in some areas. Western Corellas and Galahs have been culled – 2169 sometimes in large numbers – as part of conservation efforts to reduce nest competition 2170 with endangered Carnaby’s Black-Cockatoos (Saunders and Doley, 2017). The use of nest 2171 boxes by abundant species may not be a desirable outcome in all circumstances. There is an 2172 open season on Australian Ringnecks (Barnardius zonarius) and corellas (Cacatua spp.) 2173 across most of south-western Australia and damage permits may be issued for Australian 2174 Wood Ducks (Chenonetta jubata) in agricultural areas (Department of Biodiversity, 2175 Conservation and Attractions 2019). Any nest box programs initiated to benefit a specific 2176 species should incorporate monitoring regimes and protocols for managing use by species 2177 which managers do not wish to facilitate.

2178 Several hypotheses potentially explain the minimal use of nest boxes by boobooks. 2179 Nest boxes were deliberately placed at unoccupied sites, so it is possible that boobooks 2180 were simply absent from these areas. However, in light of the substantial dispersal capacity 2181 of boobooks in our study areas inferred from genetic and banding data presented in Chapter 2182 4, this seems does not appear to be a likely explanation. It also seems unlikely that 2183 boobooks failed to use nest boxes due to an insufficient amount of time to locate the boxes. 2184 One box was located and utilized within the first year after installation but no boxes appear 2185 to have been used by boobooks in the following breeding season. Low abundance in 2186 fragmented habitats driven by factors other than nest site availability could also potentially 2187 explain low uptake of nest boxes. Alternately, it is possible that the box design is simply not 2188 favoured by boobooks. Some preference that is not currently understood could be at work. 2189 Subtle aspects of nest box construction can impact nest box use (Lambrechts et al., 2012). 2190 In some instances these preferences can be strong and unexpected. For example, in 2191 American Kestrels (Falco sparverius), nest box dimensions had a strong effect on uptake 2192 (Bortolotti, 1994).

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2193 While it is possible that the nest box design may not have been ideal for boobooks, 2194 Australian boobooks, as a species, appear to be fairly plastic in their use of nest sites. They 2195 have been recorded using caves in treeless areas (Higgins, 1999) and old corvid nests 2196 (Sedgwick and Morrison, 1948) where tree hollows are not available. Boobooks are already 2197 known to use nest boxes and have been observed doing so in urban areas. Hogg and Skegg 2198 (1961) describe the successful nesting of a pair of New Zealand Moreporks – a closely 2199 related species – in a nest box adjacent to a noisy rifle range in Auckland, New Zealand. 2200 Prior to our study, two cases of boobooks using similarly-constructed nest boxes were 2201 documented specifically within the Perth metropolitan area. In one instance, boobooks 2202 successfully raised chicks (with artificial supplementation of food) in a nest box in the Perth 2203 suburb of Victoria Park (Wells, 2007). Beckingham (2012) also reported an adult boobook 2204 seen at the entrance of an artificial nest box placed in bushland near Lake Claremont and 2205 included a photo of a downy juvenile boobook observed nearby several weeks later.

2206 Another explanation for the low rate of nest box use by boobooks is that boobooks 2207 are not limited by nest hollow availability in either urban or agricultural bushland fragments. 2208 Despite assertion in previous literature that Australian boobooks are insectivorous, they are 2209 capable of preying on relatively large birds and mammals (Olsen, 2011a). As a consequence 2210 they are probably able to compete successfully with most other species likely to use a 2211 hollow of appropriate size. In the one occupied nest box, when chicks were temporarily 2212 removed for measurement, banding, and blood sampling, we observed the remains of 2213 Rainbow Lorikeets suggesting that the lorikeets are not likely to be effective in competing 2214 with boobooks for potential nest hollows. A similar instance was reported in which Galah 2215 nestlings were presumed to have been eaten by a boobook prior to the boobook nesting in 2216 their hollow. The remains of a male Australian Ringneck – another large hollow-nesting 2217 parrot – were subsequently found in the same nest with two boobook nestlings (Mack, 2218 1965). Other instances of boobooks taking over hollows actively used by Galahs have been 2219 reported (Schulze, 1966). Conversely, the authors have observed an instance of boobooks 2220 being evicted from a nest hollow following repeated harassment by Galahs. Common 2221 Mynas – another potential nest competitor – have also been observed harassing boobooks 2222 as they left their nest hollow but the same pair of boobooks was later photographed eating 2223 Common Mynas (Trost and Olsen, 2016). The closely related New Zealand Morepork has

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2224 also been observed successfully evicting European Starlings from a nest hollow (Hogg and 2225 Skegg, 1961). As a highly territorial generalist predator, boobooks are probably more 2226 capable than most bird species of competing successfully for nest hollows, even under 2227 circumstances where suitable hollows are limited and competition from introduced species 2228 is high. If this hypothesis is correct, supplemental provision of artificial nest hollows would 2229 not be expected to increase boobook abundance unless suitable hollows are nearly absent 2230 from an area.

2231 Care should be taken not to generalise this conclusion to all predatory species 2232 utilizing tree hollows as nest sites. Nest competition has been suggested as a possible factor 2233 contributing to the decline of Norfolk Island boobooks and severe nest hollow limitation 2234 resulting from extensive habitat loss may have played a role in their decline (Olsen, 1996). 2235 Additionally, evolution in isolation from serious competition may have reduced the capacity 2236 of this subspecies to resist introduced aggressive mainland nest hollow competitors. Hollow 2237 availability may also vary with the size of the bird and the size of the hollow required for 2238 nesting. Powerful Owls are substantially larger than boobooks, take larger prey, and are 2239 potentially even less likely to be impacted by competition for nest hollows. However, their 2240 requirement for larger nest hollows – which are often scarcer in fragmented habitats – has 2241 apparently led to failures of established pairs to breed until a suitable nest box was provided 2242 (Isaac et al., 2014a).

2243 While boobooks and other predatory birds are unlikely to be severely impacted by 2244 nest competition by most introduced bird species, they may be negatively impacted by 2245 other potential nest hollow competitors. Colonization of nest hollows of all sizes by feral 2246 honeybees has been noted to be particularly problematic in southwest Western Australia 2247 (Johnstone et al. 2015) and colonization of boobook hollows by feral bees (Johnstone and 2248 Kirkby, 2007) was specifically recorded. In some instances, owls that have apparently been 2249 stung to death by bees have been observed in hollows (Western Australian Museum, n.d.), 2250 suggesting that feral honey bee nest competition sometimes directly contributes to 2251 boobook mortality. It is suspected that, in Western Australia, the impact of feral bees on 2252 cavity nesting birds is greatest in the Wheatbelt where canola crops prompt more frequent 2253 swarming (Johnstone and Kirkby 2007). Nest boxes used in our study incorporated several 2254 features intended to deter use by feral honeybees and our results may not be

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2255 representative of honeybee competition rates for all nest box designs or in natural hollows. 2256 Anecdotally, we observed honeybees using several nest boxes unrelated to our study in the 2257 Perth metropolitan area. We encourage authors of future studies involving nest boxes to 2258 carefully report on all aspects of nest box design as this is an important and frequently 2259 overlooked factor impacting life history parameters of animals using the boxes (Lambrechts 2260 et al., 2012).

2261 Conclusion 2262 While artificial nest boxes are important tools in wildlife research and conservation, 2263 our study indicates that their use is not a panacea for every situation where hollow nesting 2264 species are in need of conservation management. In an era of heavily constrained 2265 conservation budgets, ineffective nest boxes intended to improve abundance of 2266 conservation-dependant species may divert valuable funds from more effective uses. 2267 Furthermore, poor design or application in inappropriate circumstances may lead to 2268 unintended negative outcomes for native biodiversity or unanticipated bias in scientific 2269 studies.

2270 Acknowledgments

2271 This project was supported financially by The Holsworth Wildlife Research 2272 Endowment via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird 2273 Research Award, and the Edith Cowan University School of Science Postgraduate Student 2274 Support Award. We thank the Western Australia Department of Biodiversity, Conservation, 2275 and Attractions, and the many city councils and private landowners who provided access to 2276 the sites involved in this project. This research was made possible by the generous 2277 assistance of dozens of volunteers who assisted in boobook surveys and nest monitoring. 2278 We especially thank Dr. Cheryl Lohr for providing valuable assistance in statistical analysis.

2279

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2280 Chapter 6 Toxoplasma gondii seropositivity across urban and

2281 agricultural landscapes in an Australian owl

2282

2283 Lohr, M. T., C. A. Lohr, A. H. Burbidge, and R. A. Davis. Toxoplasma gondii seropositivity 2284 across urban and agicultural landscapes in an Australian owl. Veterinary Parasitology. 2285 (In Preparation).

2286

2287 Abstract

2288 Toxoplasma gondii is an apicomplexan parasite with a wide host range and 2289 cosmopolitan distribution. House cats (Felis catus) and other members of the family Felidae 2290 are the definitive hosts for T. gondii. Members of the family Felidae were absent from 2291 Australia until house cats were brought to the continent by European explorers and 2292 colonists and the lack of evolutionary history with T. gondii has been hypothesized to leave 2293 native Australian fauna more susceptible to the negative impacts of infection. As a 2294 consequence, understanding the factors that drive differences in environmental prevalence 2295 of T. gondii may inform conservation strategies for vulnerable Australian wildlife. As cat 2296 abundance has been documented to vary with landscape composition, we hypothesized 2297 that T. gondii infection would be more prevalent in urban and agricultural landscapes than 2298 landscapes dominated by intact bushland. The Australian Boobook (Ninox boobook) was 2299 used as a test species because it has been suggested that non-migratory owls may be useful 2300 indicators of ecosystem wide T. gondii contamination. We used modified agglutination tests 2301 to determine seropositivity in serum and meat juice samples from boobooks across 2302 landscapes dominated by urban/periurban development, agriculture and intact bushland. 2303 We also examined correlations between T. gondii seropositivity and other factors like age, 2304 season, injury status, and exposure to environmental pollutants which could impact 2305 likelihood of infection. Moderately low levels of seropositivity were detected across all 2306 samples. We believe that this is the first published instance of T. gondii seropositivity in a 2307 wild predatory bird in Australia. Most risk factors previously implicated in increased risk of 2308 T. gondii infection did not show significant correlations with observed seropositivity in 2309 boobooks. However, the season in which the sample was collected did correlate

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2310 significantly with seropositivity. We suggest that seasonally dependant environmental 2311 factors which influence oocyst viability may obscure any relationship between landscape 2312 type and latent T. gondii infection rates in boobooks.

2313 Introduction

2314 Parasites and pathogens are increasingly implicated as a contributing factor in the 2315 declines of wildlife across the globe but a lack of baseline data has complicated efforts to 2316 understand the impacts of specific organisms (Pacioni et al. 2015). Well-documented severe 2317 impacts of parasitic organisms include worldwide declines in amphibian populations due to 2318 chytrid fungus (Batrachochytrium dendrobatidis) infection (Houlahan et al. 2000), ongoing 2319 reductions in North American bat populations as a result of white nose syndrome caused by 2320 the fungus Geomyces destructans (Foley et al. 2011), and the extinction or decline of most 2321 Hawaiian honeycreeper species due to avian malaria (Plasmodium relictum) (Warner, 1968). 2322 While introducing novel parasites to immunologically naïve hosts can have potentially 2323 devastating consequences, the impacts of parasites on their hosts are often more subtle 2324 (Pacioni et al. 2015). Similarly, while managing parasites and pathogens may be the key to 2325 some conservation efforts, there are very few studies on avian parasite ecology in Australia 2326 (Delgado-V. and French, 2012; Ford et al., 2001) with authors noting that “Virtually nothing 2327 is known about the effect of disease and parasites on Australian birds” (Ford et al. 2001).

2328 If parasitism is a threatening process for Australian birds, it may be exacerbated by 2329 anthropogenic land uses, sampling methodologies, or other threatening processes such as 2330 anticoagulant rodenticides (ARs) (Lemus et al., 2011; Riley et al., 2007; Serieys et al., 2018). 2331 To fully understand the implications of parasites and pathogens on avian conservation 2332 efforts it is necessary to examine patterns of parasitism across multiple habitat types. For 2333 example, Cooper’s Hawks (Accipiter cooperi) appeared to preferentially inhabit urban areas 2334 of Tucson Arizona (Battin, 2004). Urban environments tend to maintain high densities of 2335 prey species which may serve as a cue for habitat selection (Isaac et al. 2014). However, 2336 Boal & Mannan (1999) observed higher rates of nest failure in Cooper’s Hawks in urban 2337 areas than in periurban areas, due to nestlings being killed by trichomoniasis. This disease is 2338 caused by a protozoan vectored by feral pigeons which are abundant in urban areas and 2339 made up a much higher proportion of the Cooper’s Hawks’ diets in urban areas (Boal and 2340 Mannan, 1999). Hence, Cooper’s Hawks were being drawn out of less anthropogenically-

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2341 altered habitats into areas with higher prey abundance and higher risk of parasitic infection 2342 which reduced their fecundity to unsustainable levels. Documentation of demographic 2343 parameters within the city that did not explain the population’s stability indicates that the 2344 population was sustained by migration from outside urban areas and that parasites in the 2345 urban area had created an ecological trap (Battin, 2004). Similarly, birds in urban areas in 2346 Brazil were found to have higher infection rates of haemosporidian blood parasites than 2347 birds in intact natural landscapes (Belo et al., 2011). Reduced parasitism in urban areas has 2348 also been observed in some bird species but the direction of the trend is probably 2349 dependant on the type of parasite and its mode of transmission (Delgado-V. and French, 2350 2012; Suri et al., 2016). Evaluating the impact of landscape-level human land use practices 2351 on parasite prevalence will be increasingly important to the conservation of native fauna as 2352 the area of land subject to urban and agricultural development increases.

2353 Effects of Toxoplasma gondii on Humans and Wildlife 2354 Toxoplasma gondii is a parasitic protozoan capable of infecting a wide taxonomic 2355 range of warm-blooded vertebrates. Felids are its only known definitive hosts (Miller et al. 2356 1972). T. gondii causes both acute and latent toxoplasmosis (Remington and Cavanaugh, 2357 1965). It is known for its capacity for manipulation of host behaviour and increasing 2358 susceptibility to predation by cats by increasing dopamine metabolism in the brain of 2359 infected secondary hosts (Prandovszky et al. 2011). In humans, acute toxoplasmosis can 2360 cause severe illness in newborns and immunocompromised individuals and can cause 2361 spontaneous abortion and foetal abnormalities (Wolf et al. 1939). However, consensus is 2362 emerging among medical professionals that latent toxoplasmosis is not benign. It has been 2363 implicated as a risk factor in a number of serious health problems including epilepsy 2364 (Ngoungou et al. 2015), generalized anxiety disorder (Markovitz et al. 2015), schizophrenia 2365 (Torrey et al. 2007), impaired reaction time (Havlícek et al. 2001), car accidents (Flegr et al. 2366 2002), obsessive compulsive disorder (Miman et al. 2010b), Parkinson’s disease (Miman et 2367 al. 2010a), Alzheimer’s disease (Kusbeci et al. 2011), Down Syndrome (Prandota, 2010), and 2368 suicide attempts (Arling et al. 2009). Worldwide correlations with other diseases were 2369 examined by Flegr et al. (2014). It has even been proposed that T. gondii may have a 2370 worldwide impact on human culture by subtly altering the neurochemistry of substantial 2371 proportions of the global population (Lafferty, 2006). Because roughly one third of the

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2372 worldwide human population is believed to be infected with T. gondii, there is growing 2373 concern over the impacts of this organism on global human health.

2374 T. gondii infection is also a concern for native Australian wildlife. Lack of a felid 2375 definitive host prior to the arrival of Europeans suggests that native wildlife do not share a 2376 long evolutionary history with T. gondii. Acute toxoplasmosis has been observed to be 2377 lethal in a wide variety of Australian native marsupial species and can cause blindness, 2378 lethargy, respiratory and digestive problems, and decreased coordination (Patton et al. 2379 1986; Canfield et al. 1990). T. gondii infection likely increases the probability of predation in 2380 some native mammals (Obendorf et al. 1996) and has been implicated in the decline of 2381 some wild marsupial populations but the extent of its impact is not well understood 2382 (Freeland, 1994). Three macropod species were observed to be infected with multiple 2383 strains of T. gondii (Pan et al. 2012) and cat predation coupled with T. gondii infection may 2384 have played a role in the observed local decline of another macropod species (Fancourt, 2385 2014). Death by acute toxoplasmosis has also been observed in ten species of native 2386 Australian birds held in captivity (Hartley and Dubey, 1991) including one penguin which had 2387 only been in captivity for a few days (Mason et al., 1991) but prevalence of T. gondii 2388 infection has not been quantified in wild bird populations and has not been examined at a 2389 landscape level.

2390 Predatory Birds and Toxoplasma gondii Infection 2391 Toxoplasma infection has been observed in owls and other raptors in the wild in 2392 North America and Europe (Kirkpatrick et al. 1990; Lindsay et al. 1993; Dubey et al. 2010; 2393 Lopes et al. 2011; Yu et al. 2013), with some species documented to have seroprevalence 2394 rates of nearly 80% (Aubert et al. 2008). Prevalence of T. gondii infection in wildlife is a good 2395 indicator of environmental contamination by oocysts and is useful in assessing risk to 2396 human health (Dubey and Jones, 2008). Similarly, T. gondii infection is more likely to be 2397 detected in predators which typically have higher rates of seroprevalence than omnivores 2398 and herbivores as a result of greater risk of ingesting infected animals over their lifetime 2399 (Hejlícek et al., 1997; Hollings et al., 2013). Birds in particular are preferable as 2400 environmental bio-monitors for T. gondii because vertical transmission (direct congenital 2401 transmission from adult to offspring) has not been observed in Australian birds, in contrast 2402 to marsupials (Parameswaran et al. 2009). Vertical transmission in a population could make

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2403 seroprevalence rates a less useful index of overall environmental contamination. Vertical 2404 transmission is unlikely in birds due to the extremely low incidence of T. gondii in eggs 2405 (Dubey, 2010). Many raptor species prey primarily on small birds and mammals that are 2406 frequently intermediate hosts of T. gondii and, as such, raptor seroprevalence rates have 2407 the potential to offer valuable insights into environmental prevalence of T. gondii (Love et 2408 al. 2016).

2409 At present, seroprevalence rates in wild Australian raptors are unknown and within- 2410 species differences across different land-use categories have not been studied in any raptor 2411 species worldwide. Non-migratory owl species, such as Australian Boobooks (Ninox 2412 boobook), have been specifically identified as useful indicator species (Gazzonis et al., 2018) 2413 for assessing differences in environmental T. gondii oocyst contamination on a landscape 2414 scale. T. gondii is known to infect owls (Dubey et al., 1992) but direct mortality and 2415 observable illness resulting from infection are extremely uncommon (Mikaelian et al., 1997). 2416 Sub-lethal effects of T. gondii infection on owls are largely unknown but most carnivorous 2417 birds are assumed not to be affected by acute clinical toxoplasmosis (Dubey et al., 2010; 2418 Love et al., 2016). The ability to become infected without obvious signs of increased 2419 mortality rates is a desirable attribute in effective bio-monitors. Boobooks are an ideal 2420 species for monitoring landscape-level T. gondii prevalence because they are found in a 2421 wide range of habitat types across Australia – including those that have been substantially 2422 altered by humans.

2423 Aims 2424 We sought to determine whether correlations exist between different types of 2425 human land use and T. gondii infection rates in raptors. Domestic cat density has been 2426 observed to correlate strongly with housing density (Sims et al. 2008) and spatial correlation 2427 between T. gondii seropositivity and both human habitation and cat density has been noted 2428 in carnivorous wildlife (Barros et al., 2018; Hollings et al., 2013). To identify the relative 2429 importance of landscape type in risk of T. gondii infection, we also assessed other factors 2430 associated with increases in toxoplasma seropositivity in wildlife including age (Cabezón et 2431 al., 2011; Lindsay et al., 1993; Lopes et al., 2011), injury status (Hollings et al., 2013), and 2432 sampling during seasons with more favourable conditions for T. gondii oocyst survival 2433 (Simon et al., 2018). We also explored potential associations between T. gondii

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2434 seropositivity and anticoagulant rodenticides (ARs) because an emerging body of research 2435 has shown correlations between ARs and parasites and infectious diseases (Lemus et al., 2436 2011; Riley et al., 2007; Serieys et al., 2015; Vidal et al., 2009).

2437 Consequently, we expected to observe the highest rate of boobook seropositivity in 2438 urban and periurban areas, where both cats and commensal rodents exist in elevated 2439 densities as a result of human activities. We hypothesized that:

2440 1) Seroprevalence in agricultural landscapes would be intermediate between 2441 seroprevalence observed in primarily urban/periurban landscapes and landscapes 2442 dominated by native bushland.

2443 2) T. gondii seropositivity will be higher in individuals which are older, sampled 2444 during wetter seasons, and in the dead or injured category due to increased reaction time 2445 associated with infection potentially increasing the risk of collisions with windows and 2446 motor vehicles.

2447 3) T. gondii seropositivity will be higher in birds that have detectable levels of 2448 anticoagulant rodenticides (ARs).

2449 Methods

2450 Sample Collection 2451 We used several methodologies to collect boobook blood and tissue samples across 2452 a variety of habitat types present in Western Australia, with an active focus on procuring 2453 samples in the Perth Metropolitan Area, areas of intensive agriculture within an 2454 approximate 60km radius of the town of Kellerberrin in the central wheatbelt approximately 2455 200 km east of Perth, and intact forested areas of the Perth Hills between the two types of 2456 fragmented landscape. During occupancy surveys for another study (Chapter 5), boobooks 2457 were located at night across all three habitat types using recorded boobook calls broadcast 2458 on a portable speaker. Additionally, boobooks roosting during the day were located with 2459 the assistance of volunteers and were also opportunistically included in the study. Wild 2460 boobooks were captured using a noose pole similar to methodology used to capture 2461 boobooks elsewhere (Olsen et al., 2011). After banding and basic biometric measurements, 2462 boobooks were assigned to age classes of one year or less ('hatch year') or greater than one

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2463 year ('after hatch year') based on the presence of juvenile down and by examination of 2464 fluorescence patterns in the undersides of the remiges under ultraviolet light (Weidensaul 2465 et al., 2011). In some instances, age class could not be assigned because of degradation of 2466 porphyrins by exposure to sunlight. Following these measurements, 0.5 ml of blood was 2467 taken from the jugular vein of all boobooks. Live boobooks held by wildlife rehabilitators 2468 were also sampled if the bird was sufficiently healthy for blood sampling. Blood was taken 2469 from the right jugular vein to reduce handling time and risk of hematoma relative to 2470 sampling from the brachial vein (Owen, 2011). Blood taken from these boobooks was 2471 allowed to coagulate for approximately 24 hours. Samples were then centrifuged for 10 2472 minutes at 13,200 RPM to produce serum. Serum was stored at -20°C until testing.

2473 Boobooks found dead by volunteers or euthanized by wildlife rehabilitators were 2474 opportunistically sampled as well (see methods in Lohr, 2018). Heart and breast muscle 2475 tissue were removed from dead boobooks which were not in an advanced state of 2476 decomposition. These tissues were placed in a plastic bag, and stored frozen at -20°C. Prior 2477 to testing, the specimens were thawed and 0.5ml of resulting fluids (hereafter “meat juice”) 2478 was removed from the bag using a syringe. Meat juice samples were then centrifuged for 2479 10 minutes at 13,200 RPM and the supernatant was removed and stored at 4°C until it was 2480 used for testing within 24 hours of tissue thawing.

2481 Serological Testing 2482 Serum and meat juice were both tested using a commercially available modified 2483 agglutination test (Toxo-Screen DA, BioMerieux, France). Modified agglutination tests 2484 (MATs) are the preferred serologic tests used in detecting chronic toxoplasma infection in 2485 wild birds because they are sensitive, specific, do not require special equipment, and appear 2486 to work well across all avian species tested (Dubey, 2002). Testing was conducted according 2487 to the instructions included with the commercial kit. The only variation from the testing kit 2488 instructions was that we used serum dilutions of 1:25 and 1:400 and meat juice dilutions of 2489 1:4 and 1:64. This synchronized maximum concentrations with those used in previous 2490 testing of similar matrices (Cabezón et al., 2011; El-Massry et al., 2000) and (Richomme et 2491 al., 2010), respectively) using the same commercially available testing kit and maintained 2492 equal dilution ratios between the two matrix types we tested. A higher concentration of 2493 meat juice was used in testing because meat juice contains lower concentrations of

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2494 Toxoplasma antibodies than serum (Richomme et al., 2010). In accordance with previous 2495 literature, we considered positive results in samples diluted at a ratio of 1:25 as indicative of 2496 latent Toxoplasma infection in serum samples (Aubert et al., 2008; El-Massry et al., 2000) 2497 and used a threshold of 1:4 for determining seropositivity in meat juice samples (Richomme 2498 et al., 2010). Positive and negative controls provided with the kit were included in each 2499 testing plate. A total of 130 individuals were tested. Of these, 61 were tested using only 2500 serum, 61 were tested using only meat juice and eight were tested using both serum and 2501 meat juice.

2502 In eight instances, both serum and meat juice were available for the same individual 2503 birds due to either subsequent euthanasia of boobooks held in care which failed to recover 2504 sufficiently to allow release or banded boobooks being handed in by members of the public 2505 after being found dead.

2506

2507 Statistical Analysis 2508 We used RStudio 1.1.383 (RStudio, Inc., Boston,MA, USA) to conduct Fisher's exact 2509 tests in order to examine correlations between T. gondii seropositivity and a number of 2510 potentially relevant environmental and demographic variables because in all tests the 2511 number of observations in at least one category was ≤ 5 (Gazzonis et al., 2018). Tested 2512 variables included landscape type (agriculture, bushland, urban/periurban), the age of the 2513 boobook sampled (hatch year or after hatch year), season in which the sample was collected 2514 (winter, spring, summer, autumn), the status of the boobook when sampled (wild or 2515 compromised (in care or dead)). Sample sizes varied slightly between the individual 2516 statistical tests because, in some cases, volunteers provided incomplete collection 2517 information regarding collection date and location or because we were unable to accurately 2518 determine the age of the boobook. This necessitated the exclusion of some deceased 2519 individuals from particular statistical tests.

2520 A subset of deceased boobooks (N=65) tested for AR residues in a previous study 2521 (Lohr, 2018) were used to test for correlations between T. gondii seropositivity and total 2522 concentrations of ARs in liver tissue. Individuals were assigned to four categories of AR 2523 exposure based on biologically relevant thresholds. The lowest category included all

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2524 samples with total AR concentration below 0.01 mg/kg because it is the limit of 2525 quantification for most of the ARs tested (Lohr, 2018). Additional thresholds of 0.10 mg/kg 2526 and 0.50 mg/kg were also used to delineate the remaining three categories because 0.10 2527 mg/kg is regularly used as the lower limit for potential toxic effects in raptors (Albert et al., 2528 2010; Christensen et al., 2012; Langford et al., 2013; Ruiz-Suárez et al., 2014; Shore et al., 2529 2016; Stansley et al., 2014; Walker et al., 2011, 2008) and liver concentrations of 0.50 mg/kg 2530 are likely to be lethal in most birds (Dowding et al., 1999). We used Fisher’s exact test to 2531 determine if toxoplasma seropositivity was associated with AR exposure.

2532 Results 2533 In the eight boobooks with both serum and blood samples available, six were 2534 negative in both samples and two appeared to seroconvert and had negative serum samples 2535 but positive meat juice samples. Across all 130 boobooks sampled, 13.1% were seropositive 2536 for T. gondii in at least one sample. Seropositivity was more prevalent in meat juice samples 2537 (18.0% n = 61) than in serum samples (6.6% n = 61) but did not differ significantly between 2538 the two matrix types sampled (P=0.096). Consequently, for analyses other than direct 2539 comparisons between the serum and meat juice seropositivity and for comparisons 2540 involving AR exposure which was only testable in dead boobooks, the data from both 2541 matrices were pooled and boobooks testing positive in either matrix were treated as 2542 positive samples. The only factor which significantly correlated with T. gondii seropositivity 2543 was the season in which the sample was collected (p = 0.024) (Error! Reference source not 2544 ound.). While subsequent pairwise comparisons between seropositivity by season were not 2545 significant, seropositivity rates were numerically higher in autumn and winter relative to 2546 spring and summer (Error! Reference source not found.) and the difference between 2547 ummer and autumn seropositivity rates was marginally non-significant (p = 0.099). Overall 2548 anticoagulant rodenticide exposure did not show significant associations with T. gondii 2549 seropositivity but seroprevalence was numerically lower in boobooks with total liver 2550 concentrations of ≤ 0.01mg/kg (13.0%) than in the three higher categories (23.7% to 25%) 2551 (Figure 6.2).

2552 Table 6.1 Factors associated with Toxoplasma gondii seroprevalence in Australian Boobooks (Ninox boobook) in Western 2553 Australia.

Variable Category Positive/ examined Seroprevalence (%) p-value Testing matrix serum 4/61 6.6 0.096

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meat juice 11/61 18.0 Age < 1 year 12/88 13.6 1.000 ≥ 1 year 5/36 13.9 Landscape type Agriculture 3/17 17.6 0.306 Bushland 0/14 0.0 Urban/Periurban 14/90 15.6 Injury status Wild 4/42 9.5 0.580 In care/dead 13/88 14.8 Season Summer 3/60 5.0 *0.024 Autumn 8/35 22.9 Winter 3/12 25.0 Spring 2/22 9.1 AR exposure 0-0.01 mg/kg 3/23 13.0 0.759 0.01-0.10 mg/kg 2/8 25.0 0.10-0.50 mg/kg 5/22 22.7 >0.50 mg/kg 3/12 25.0 2554

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2555

2556 Figure 6.1 Seasonal Toxoplasma gondii seroprevalence in Australian Boobooks (Ninox boobook) in Western Australia. Width 2557 of the bars is representative of sample size.

2558

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2559

2560 Figure 6.2 Toxoplasma gondii seroprevalence in meat juice from deceased Australian Boobooks (Ninox boobook) in Western 2561 Australia in four different categories of anticoagulant rodenticide exposure (A= ≤ 0.01 mg/kg, B=0.01 mg/kg – 0.10 mg/kg, 2562 C 0.10 mg/kg - 0.50mg/kg, D ≥ 0.50mg/kg) Width of the bars is representative of sample size.

2563 Discussion

2564 Apparent seroconversion in the two individuals which tested negative in serum 2565 samples but positive in meat juice samples may be an artefact of the MAT test used. False 2566 negative results can be obtained during acute stages of T. gondii infection because the test 2567 is only sensitive to IgG antibodies, and not IgM antibodies which are present at the onset of 2568 infection (Sroka et al., 2008). It is entirely plausible that the two boobooks were 2569 experiencing active infections when initially sampled but their infections would only be 2570 detected by the subsequent meat juice sampling. We believe that this explanation, in 2571 combination with the lack of a significant difference in seropositivity rates between serum 2572 and meat juice samples justifies our decision to combine data from both matrix types in the 2573 other analyses.

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2574 Studies using the same commercially available modified agglutination test on other 2575 continents have found varying rates of seropositivity across multiple raptor species: 34.5% 2576 in the south-eastern USA (n=281) (Love et al., 2016), 25.7% in Taiwan (n=206) (Chen et al., 2577 2015), 35.8% in France (n=53) (Aubert et al., 2008), 50.0% in Portugal (Lopes et al., 2011), 2578 and 10.7% in Italy (n=93) (Gazzonis et al., 2018). While at the lower end of the scale, our 2579 results (13.1%) were within the ranges previously reported in studies of predatory birds. 2580 Interestingly, overall seropositivity was nearly identical to the rate of 13.0% reported for a 2581 native marsupial carnivore (chuditch, Dasyurus geoffroii), in Julimar Valley (Parameswaran, 2582 2008), an area of continuous bushland adjacent to our sites in the Perth Hills.

2583 Four potential explanations exist for the relatively low seropositivity rates we 2584 observed. Australian boobooks have not previously been evaluated using this test and it is 2585 possible that species-specific factors may have led to false negative results. We view this 2586 scenario as unlikely because, while false negatives using MAT are common in some species – 2587 particularly in dogs (Liu et al., 2015) – this test has been used successfully to detect 2588 toxoplasma seropositivity in a wide variety of other predatory bird species (Chen et al., 2589 2015; Gazzonis et al., 2018; Lopes et al., 2011).

2590 It is also unlikely that our use of meat juice in addition to serum would have reduced 2591 detections relative to other studies. A study directly examining detection of T. gondii 2592 antibodies in meat juice did not find reduced detectability or degradation of antibodies in 2593 response to repeated freezing and thawing of meat (Mecca et al., 2011). If anything, the 2594 use of this methodology should have increased seropositivity detection in our study relative 2595 to other studies which tested only serum. The numerically but not significantly higher 2596 detection rate of T. gondii antibodies in meat juice samples is consistent with this 2597 hypothesis.

2598 The diet and trophic position of boobooks may also provide some explanation for 2599 the relatively low seropositivity rates seen in boobooks. Australian Boobooks are medium- 2600 sized owls (Olsen, 2011a) and consume a variety of invertebrate and vertebrate prey (Trost 2601 et al., 2008). A study on T. gondii seropositivity in wild birds in Spain found seropositivity 2602 rates ranging from 0% to 25% in six small and medium sized owl species (Cabezón et al., 2603 2011). However, the same study detected T. gondii antibodies in 68% of all Eurasian Eagle-

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2604 owls (Bubo bubo) (Cabezón et al., 2011). This species is substantially larger and occupies a 2605 higher trophic level than the other owl species tested. In the context of this research, the 2606 seropositivity of boobooks we observed is typical of an owl species of its size and diet.

2607 Landscape Type 2608 The relatively warm and dry climate in our study areas may explain our observation 2609 of lower seropositivity rates than in most other areas of the world where raptors have been 2610 sampled. Worldwide, toxoplasma prevalence is lowest in hot arid areas, presumably due to 2611 shorter duration of oocyst viability under hot dry conditions (Meerburg and Kijlstra, 2009). 2612 In a study examining habitat impacts on T. gondii seropositivity in wild rabbits, seropositivity 2613 was substantially higher in habitats with more shade and humidity (Almería et al., 2004).

2614 This pattern may explain the lack of significant difference observed in seropositivity 2615 between landscape types. Counter-intuitively, clearing of land for urban and agricultural 2616 uses could lead to a reduction in T. gondii seroprevalence despite potential increases in cat 2617 abundance if the reduction in vegetative cover results in an increase in soil temperature and 2618 decrease in soil moisture, leading to inhibition of T. gondii oocyst viability. Future work 2619 examining T. gondii seroprevalence in a single intermediate host species across paired 2620 habitat types in areas with substantially different rainfall levels would be useful in 2621 determining relative contributions of cat abundance and soil moisture to seropositivity in 2622 intermediate hosts.

2623 Age 2624 We were surprised that no difference in seropositivity was detected between age 2625 classes. Some studies have found that T. gondii detection increases with age in wild 2626 predatory birds (Cabezón et al., 2011; Lindsay et al., 1993; Lopes et al., 2011) which is in 2627 keeping with the hypothesized lifelong persistence of the parasite after infection. However, 2628 other studies of predatory birds (Gazzonis et al., 2018) and wild rabbits (Oryctolagus 2629 cuniculus) (Almería et al., 2004) have failed to detect a difference in seropositivity between 2630 different age classes but did not address why no correlation was detected. It is possible that 2631 our grouping of boobooks into two coarse age classes of < one year and ≥ one year obscured 2632 longer-term trends in seropositivity. Because boobooks are relatively long lived – one was 2633 re-sighted in the field alive nearly 16 years after it was originally banded (Commonwealth of

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2634 Australia, 2015) – they are potentially at risk from the long-term effects of latent T. gondii 2635 infection similar to those observed in humans. Impaired reaction time resulting from T. 2636 gondii infection is cumulative in humans and increases with the duration of latent infection 2637 (Havlícek et al. 2001). Similar effects in long-lived wildlife could predispose individuals to 2638 greater risk of predation, accident, or vehicular collision. Use of predatory bird species with 2639 a greater number of more easily-identifiable age categories (such as Wedge-tailed Eagles 2640 (Aquila audax)) could help to resolve questions relating to both changes in seropositivity 2641 between age classes and whether cumulative impacts of latent toxoplasmosis are 2642 problematic for predatory birds.

2643 Injury Status 2644 The lack of significant difference in seropositivity between boobooks found dead or 2645 held by wildlife carers and those captured in the wild was also unexpected and runs 2646 contrary to observations by Hollings et al. (2013) in Tasmanian pademelons (Thylogale 2647 billardierii) shot for pest control purposes and pademelons killed by motor vehicle collisions. 2648 It is unlikely that this is a consequence of our testing of multiple matrix types, as 2649 seropositivity was numerically – though not significantly – higher in meat juice samples from 2650 deceased boobooks. It seems more likely that our inclusion of boobooks which were killed 2651 or disabled by a wide variety of causes may have obscured any potential effect specific to 2652 motor vehicle collisions. In humans, reduced concentration time and increased reaction 2653 time were proposed as the mechanisms by which T. gondii seropositivity increased rates of 2654 car accident (Flegr et al., 2002). It is unlikely that these potential causative factors are 2655 relevant to all the causes of mortality or injury associated with the boobooks in our study. 2656 Unfortunately, uncertainty over proximate causes of death or injury in the boobooks we 2657 tested precluded direct testing of a more specific relationship with seropositivity.

2658 Season 2659 Several potentially interacting biological factors could plausibly explain the higher 2660 seropositivity rates of boobook samples obtained in autumn and winter. Boobooks 2661 consume a higher proportion of mammals and birds in winter relative to other times of year 2662 when insects make up a larger percentage of their diet (Trost et al., 2008). Additionally, 2663 temperature and rainfall patterns in autumn and winter in southwest Western Australia are 2664 more conducive to T. gondii oocyst viability and, as a consequence, infection rates in prey

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2665 species may be higher at this time of year. A similar explanation was given for observations 2666 of higher T. gondii seroconversion rates observed in house cats in autumn and winter 2667 relative to spring and summer (Simon et al., 2018). Both factors may increase the risk of 2668 boobooks consuming prey with tissue cysts containing T. gondii bradyzoites and subsequent 2669 infection in winter. Boobooks with recent infections may also have been easier to capture 2670 or more likely to die and be injured and consequently be sampled by our study, increasing 2671 the detected seropositivity of individuals sampled at this time of year.

2672 Anticoagulant Rodenticide Exposure 2673 Alternately, exposure to anticoagulant rodenticides may have contributed to 2674 increased seropositivity of samples obtained in winter. Significantly higher liver 2675 concentrations of anticoagulant rodenticides have been observed in boobooks in the Perth 2676 metropolitan area in winter relative to spring and summer (Lohr, 2018). Additionally, while 2677 the Fisher’s exact test failed to detect a significant difference between rodenticide exposure 2678 categories, the seroprevalence of boobooks in the lowest exposure category with 2679 insubstantial amounts of rodenticide was numerically lower than the three categories with 2680 clinically relevant AR exposure (≥0.01 mg/kg) (Figure 6.2). Sub-lethal exposure to 2681 anticoagulant rodenticides has been found to correlate with immune dysfunction in bobcats 2682 (Lynx rufus) (Serieys et al., 2018) and has been hypothesized as the explanation for an 2683 observed correlation between anticoagulant rodenticides and notoedric mange (Riley et al., 2684 2007). These correlations are to some degree called into question by a study on domestic 2685 cats (Felis catus) which did not find a substantial link between anticoagulant rodenticides 2686 and immune dysfunction (Kopanke et al., 2018). However, even if immunosuppression is 2687 not the mechanism by which AR exposure facilitates hyper-parasitism, similar increases in 2688 pathogen and parasite load correlated with AR exposure have also been documented in 2689 Great Bustards (Otis tarda) exposed to the AR chlorophacinone (Lemus et al., 2011).

2690 If AR exposure facilitated reactivation of latent toxoplasmosis, this could explain the 2691 increase in seroprevalence detected in winter and autumn. Alternately, it is possible that a 2692 synergistic interaction between AR exposure and T. gondii infection increased the 2693 probability of the boobooks dying and entering this study to be tested. A synergistic effect 2694 on probability of mortality involving the AR chlorophacinone and the pathogen Francisella 2695 tularensis has been suggested in common voles (Microtus arvalis) (Vidal et al., 2009).

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2696 Another hypothesis potentially explaining the possible correlation between AR exposure 2697 and seropositivity is that boobooks could simply be exposed to both T. gondii and ARs at 2698 higher rates in winter months as a consequence of rodents making up a higher proportion of 2699 their diet at this time of year. These hypotheses are not mutually exclusive and would be 2700 difficult to distinguish without experimental study of this dynamic in a laboratory setting.

2701 While boobooks showed few significant trends in T. gondii seropositivity, this may be 2702 primarily an issue of low statistical power to detect such trends caused by relatively low 2703 sample sizes. This is a common problem when studying cryptic, nocturnal, carnivores which 2704 occurr at low densities and are difficult to capture. However, seasonal differences in 2705 seropositivity suggest that conditions influencing oocyst viability may be a more important 2706 determinant of exposure risk than the factors we examined directly. Future work evaluating 2707 the utility of boobooks and other raptors as bioindicators of environmental T. gondii 2708 contamination should examine seropositivity rates across temperature and rainfall 2709 gradients. The use of boobooks as bioindicators could help identify important landscape- 2710 level drivers of T. gondii prevalence and has the potential to inform management actions 2711 and translocation efforts intended to benefit susceptible native mammals.

2712 Acknowledgments

2713 This project was supported financially by The Holsworth Wildlife Research Endowment 2714 via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird Research 2715 Award, and the Edith Cowan University School of Science Postgraduate Student Support 2716 Award. We thank Annette Koenders, Adriana Botero, and Louise Pallant for advice and 2717 technical assistance in serology testing. Our research would not have been possible without 2718 contributions of samples and access to live birds provided by Kanyana Wildlife 2719 Rehabilitation, Native Animal Rescue, Native ARC, Nature Conservation Margaret River 2720 Region, Eagles Heritage Wildlife Centre, and many individual volunteers especially Simon 2721 Cherriman, Angela Febey, Amanda Payne, Stuart Payne, and Warren Goodwin.

2722

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2723 Chapter 7 Summary, Synthesis, and Management Implications

2724 I examined four distinct potential threatening processes which I predicted had the 2725 potential to vary in magnitude of impact between habitats fragmented by urban and 2726 agricultural land uses. Australian Boobooks (Ninox boobok) did not appear to be 2727 substantially negatively impacted by lack of nest hollow availability, infection with 2728 Toxoplasma gondii, or genetic isolation in either landscape type. However, I did detect 2729 considerable exposure to anticoagulant rodenticides (ARs) associated with the use of 2730 habitats containing commercial and residential development.

2731 In this chapter, I highlight the most important findings of each chapter and 2732 contextualise their relevance to their respective fields outside of the specific system I 2733 studied. I also discuss the contribution of my research to the theoretical framework in 2734 which the impacts of habitat fragmentation are typically evaluated. I then suggest specific 2735 practical implications of my findings for management actions intended to maintain or 2736 increase native biodiversity in landscapes dominated by intensive human land uses.

2737 Summary of major findings:

2738 Objective 1. Critically review literature on anticoagulant rodenticide exposure in native 2739 wildlife in Australia to clarify its role as a threatening process. 2740 My review of the literature relating to anticoagulant rodenticides in Australia 2741 revealed widespread anecdotal accounts of both primary and secondary anticoagulant 2742 rodenticide (AR) poisoning among a taxonomically diverse group of non-target wildlife. Key 2743 differences between Australia and other developed nations were noted in the regulation of 2744 ARs. Most notably, second generation anticoagulant rodenticides (SGARs) are readily 2745 available for purchase without a license in Australia, unlike in the United States and Canada. 2746 Australia is also one of only two countries to allow the use of the first generation 2747 anticoagulant rodenticide (FGAR), pindone and to allow its use in widespread repeated 2748 baiting of natural systems for control, rather than eradication, of introduced species. 2749 Additional research is recommended to evaluate this practice. The review also identified 2750 patterns in world literature relating to reptiles and rodenticides which suggest the potential 2751 for high tolerance to rodenticides in at least some reptile taxa. Further experimental testing 2752 is necessary to determine if this hypothesized resistance makes reptiles efficacious vectors

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2753 of ARs to humans and predatory wildlife. If so, rodenticide poisoning in warmer areas of the 2754 world with diverse and abundant reptile herpetofaunas, may be a greater threat to 2755 predatory wildlife than in the cool temperate regions where most AR ecotoxicology work 2756 has been conducted.

2757 Objective 2. Investigate the relationship between exposure to anticoagulant rodenticides 2758 and urban and agricultural fragmentation. 2759 Exposure to anticoagulant rodenticides (ARs) was prevalent in the boobooks tested 2760 (72.6%) and higher than typically observed in similar studies of predatory birds on other 2761 continents. The vast majority of the rodenticides detected were the more persistent second 2762 generation anticoagulant rodenticides (SGARs). AR exposure correlated positively with 2763 proximity to urban/periurban habitat at all spatial scales and negatively with use of 2764 agricultural areas and native bushland. The association between AR exposure and the 2765 proximity of boobooks to urban and suburban development (but not agricultural land uses), 2766 supports modelling which suggests that matrix type can exert strong influences on wildlife 2767 inside habitat patches (Sisk et al., 1997). The strongest correlations between AR exposure 2768 and habitat were found at the spatial scale of a boobook’s estimated home range. This 2769 suggests that predatory birds with larger home ranges may be at risk of AR exposure over a 2770 larger proportion of the landscape. Additional research on non-target AR exposure in 2771 Australia is urgently needed to determine the level of threat posed to other wildlife species, 2772 particularly carnivores and scavengers with large home ranges which are already listed as 2773 threatened (e. g. quolls (Dasyurus sp.) and Tasmanian devils (Sarcophilus harrisiii).

2774 Objective 3. Determine if urban and agricultural fragmentation influence boobook genetic 2775 structure. 2776 Boobooks did not exhibit substantial genetic structure among landscapes dominated 2777 by urban development, agricultural crops, or native bushland in between. This trend held 2778 with the inclusion of boobook samples originating across a larger geographic area including 2779 the majority of Western Australia. Banding data from my study and others demonstrated 2780 that fledgling boobooks are capable of dispersing across urban habitats for distances far 2781 greater than those between remaining bushland fragments. In combination, these findings 2782 suggest a high degree of landscape permeability and genetic connectivity in boobooks 2783 across all areas sampled. Highly mobile species have a greater probability of survival than

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2784 less mobile species in areas which have experienced habitat fragmentation (Ewers and 2785 Didham, 2006). High mobility despite fragmentation coupled with the apparent capacity to 2786 use matrix habitat in at least some circumstances likely explains the persistence of 2787 boobooks in highly fragmented landscapes, albeit at lower densities.

2788 Objective 4. Examine whether nest box supplementation increases site occupancy at 2789 unoccupied sites and whether this effect differs between urban and agricultural landscapes. 2790 Boobooks occupied fewer sites in urban and agricultural remnant bushlands than in 2791 continuous woodland. Nest box supplementation at unoccupied sites did not alter site 2792 occupancy over the duration of this study. However, one nest box in an urban bushland 2793 remnant was successfully used by a boobook. Nest hollows do not appear to be a limiting 2794 factor in the use of remnant woodlands by boobooks in either fragmented landscape type 2795 despite boobooks being obligate hollow nesters. Nest box supplementation is unlikely to be 2796 an effective tool for increasing boobook abundance in remnant woodlands but anecdotal 2797 observations of boobooks utilising nest boxes in urban areas completely devoid of native 2798 bushland suggest that nest boxes may reduce matrix hostility and increase usable space in 2799 highly-altered areas lacking remaining suitable tree hollows.

2800 Objective 5. Explore patterns of Toxoplasma gondii seropositivity in boobooks across the 2801 urban, agricultural, and natural landscapes. 2802 Toxoplasma gondii seropositivity did not vary significantly among urban, agricultural, 2803 and woodland dominated landscape types. Most other factors which other studies have 2804 found to correlate with T. gondii seropositivity (i.e. age, season, injury status, and exposure 2805 to environmental pollutants) did not show significant correlations. Failure to detect these 2806 trends may have been caused by insufficient statistical power associated with low 2807 seropositivity rates. However, higher seropositivity was observed in cooler wetter seasons. 2808 This trend could be related to environmental conditions favouring oocyst viability, greater 2809 availability of infected prey, seasonal dietary shifts toward increase proportional 2810 consumption of prey species likely to be infected, or a combination of these factors. 2811 Increased risk of boobooks being infected by T. gondii associated with increased numbers of 2812 house cats (the definitive hosts for T. gondii) in urban and agricultural landscapes may be 2813 offset by decreased viability of oocysts in soil due to increases in soil temperature and 2814 decreases in soil moisture relative to areas of natural vegetation. This chapter reports what I

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2815 believe to be the first confirmation of T. gondii seropositivity in a wild predatory bird species 2816 in Australia.

2817 Synthesis

2818 My genetic research, in combination with band return data, revealed a biological 2819 reality more complex than the initial premise underpinning my research. I had initially set 2820 out to determine whether potential threatening processes operated differently in 2821 landscapes dominated by different anthropogenic matrices (i.e. urban and agricultural land 2822 use). Habitat fragmentation defined as “the process of subdividing a continuous habitat 2823 into smaller pieces” (Andrén, 1994) clearly occurred in landscapes consisting of 2824 predominantly urban or agricultural land use examined in this study, if pre-existing natural 2825 vegetation is considered to be synonymous with “habitat”. However, the lack of observed 2826 spatial genetic structure, dispersal of boobooks across urban matrix, active use of the 2827 agricultural matrix, and observed breeding inside highly developed urban areas with no 2828 adjacent bushland all suggest that boobooks are using human-dominated land cover types 2829 as well as remnant bushlands. If “habitat” is defined as “the subset of physical 2830 environmental factors that a species requires for its survival and reproduction” (Block and 2831 Brennan, 1993) then, by this definition there has been no “habitat fragmentation”. 2832 Essentially, from the perspective of boobooks, functional reduction in available habitat may 2833 not have occurred in landscapes dominated by agriculture and urban development despite 2834 extensive conversion of natural vegetation types. The misuse of the term “habitat” to mean 2835 something akin to “vegetation association” is common in published scientific literature 2836 (Franklin et al., 2002; Hall et al., 1997). The unresolved ambiguity and continuing misuse of 2837 the term “habitat” has led to the coining of the largely synonymous term “usable space” as 2838 an “area with habitat compatible with the physical, behavioral, and physiological 2839 adaptations of [an organism] in a time-unlimited sense” (Guthery et al., 2005).

2840 In the instance of boobooks examined in this study, true fragmentation of ‘usable 2841 space’” does not appear to have occurred across all areas of urban and agricultural land use. 2842 Boobooks were observed successfully fledging chicks in an area >3km from the nearest 2843 remaining patch of native vegetation and foraging in agricultural areas >1km from the 2844 nearest bushland, tree line, or patch of native vegetation (Chapter4). These behavioural 2845 observations are not conclusive but are strongly indicative that at least urban areas

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2846 constitute ‘usable space’. Despite the apparent capacity of boobooks to successfully use 2847 urban and agricultural landscapes, the conversion of native vegetation associations to urban 2848 and agricultural land uses appears to have caused habitat degradation. This is supported by 2849 our observation of lower boobook occupancy rates in urban and agricultural bushland 2850 remnants, relative to areas of intact bushland (Chapter 5). While there are no specific areas 2851 where boobooks can be defined as verifiably extirpated, their density in the landscape 2852 appears to be substantially reduced in both of the two fragmented habitats, indicating lower 2853 habitat quality rather than the absence of usable space. Boobooks’ continued use of 2854 substantially altered landscapes is likely facilitated by the same traits which allow them to 2855 use the majority of vegetation types throughout Australia.

2856 The responses of particular species to habitat fragmentation can be impacted by 2857 species-specific traits including “trophic level, dispersal ability and degree of habitat 2858 specialisation” (Ewers and Didham, 2006). While traditional fragmentation models which 2859 assume a completely hostile matrix between islands of usable habitat are still likely to apply 2860 to species which are dietary or habitat specialists, they are probably less relevant when 2861 applied to more generalist species. Species responding to apparent fragmentation by 2862 making extensive use of resources in the matrix are often classified as “urban exploiters” 2863 (Conole and Kirkpatrick, 2011).The term “urban adapters” is often used to describe species 2864 with a lower capacity to tolerate urban development but these traits exist on a continuum 2865 (Callaghan et al., 2019). The same principle applies to species responding in a similar 2866 fashion to intensive agricultural land use. Boobooks continued presence in urban and 2867 agricultural landscapes, genetic connectivity, and observed capacity to use resources 2868 derived from urban and agricultural matrix coupled with reduced detection rates relative to 2869 areas of intact bushland suggest that they function as ‘adapters’ in urban and agricultural 2870 landscapes. Species which function as exploiters or adapters are probably inappropriate for 2871 use in modelling general impacts of habitat fragmentation because these groups 2872 disproportionally share generalist functional and morphological characteristics and are not 2873 representative of the previously existing suite of taxa present prior to extensive alteration of 2874 vegetation types (Conole and Kirkpatrick, 2011).

2875 Boobooks were initially chosen because they were present across all landscape types 2876 included in the study. Future studies specifically examining responses to habitat

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2877 fragmentation should also consider key natural history traits linked to a species’ capacity to 2878 use altered habitat types when selecting study species. When dietary and habitat 2879 generalists with broad niche breadths are used, studies are less likely to detect impacts of 2880 threatening processes which may have already caused extirpation of even closely related 2881 species. For example, although boobooks in this study did not appear to be limited by nest 2882 hollow availability (Chapter 5), the congeneric Powerful Owl (Ninox strenua) appears to be 2883 limited due to its requirement of larger nest hollows which are less available in urban areas 2884 (Isaac et al., 2014a).

2885 If, on the other hand, a strong signal of detrimental effect from a specific proposed 2886 threatening process is apparent in species which are robust to overall habitat alteration, 2887 further evaluation of that threat over the spatial area where it is likely to occur is warranted. 2888 In the context of my study, the widespread and often severe exposure of boobooks to 2889 second generation anticoagulant rodenticides meets these criteria. As discussed in Lohr 2890 (2018), species of predatory birds with more specialised diets containing a larger proportion 2891 of rodents and species with larger home ranges are likely to be at greater risk of exposure to 2892 ARs. Anecdotally, raptor species meeting this description (i.e. Wedge-tailed Eagles (Aquila 2893 audax) and Masked Owls (Tyto novaehollandiae)) are largely or completely absent from the 2894 Perth metropolitan area but present outside its margins. Preliminary testing of these 2895 species (James Pay pers. comm., Michael Lohr unpublished data) has revealed exposure 2896 patterns in line with the predictions made in Lohr (2018). The detection of this pattern in 2897 boobooks despite their apparent capacity to use highly-altered landscapes, may be possible 2898 because the threat of SGARs itself is more recent than a large proportion of the habitat 2899 alteration as these chemicals were not invented until the late 1970s and early 1980s.

2900 Alternately, the continued presence of boobooks in urban areas despite the severity 2901 of the threat may be a function of their greater abundance on the wider landscape and 2902 immigration from adjacent unaffected areas masking the effects of suboptimal population 2903 parameters within urban areas. Under this circumstance, it is possible that the ability of 2904 boobooks to utilise highly-altered urban areas may pose a greater risk than if they actually 2905 experienced true habitat fragmentation with its concomitant loss of “usable space” and 2906 landscape permeability. Somewhat counterintuitively, a review of landscape-level 2907 fragmentation studies found documentation of positive effects of habitat fragmentation

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2908 including “spreading of risk, reduced competition, and stabilization of predator-prey 2909 interactions” (Fahrig et al., 2019). The tendency of boobooks to use and capacity of 2910 boobooks to move through highly-altered habitats was demonstrated in Chapter 4. If 2911 demographic parameters in urban or agricultural landscapes are such that local populations 2912 are not numerically self-supporting and fragmentation does not pose an impediment to 2913 movement, these areas, at best, may be sinks where populations in altered habitats are 2914 subsidised by dispersal from areas of less-degraded habitat.

2915 If, however, maladaptive selection cues lead boobooks to preferentially select highly 2916 altered habitats, these areas may actually be ecological traps and could lead to declines in 2917 adjacent areas of objectively better quality habitat. A similar dynamic has been observed in 2918 Powerful Owls. Urban habitat does not appear to be an effective barrier to dispersal and 2919 fledglings have been observed travelling distances up to 18 km across urban habitat (Hogan 2920 and Cooke, 2010). In this species, habitat selection appears to be driven by availability of 2921 mammalian prey which are common in urban areas but a lack of available nesting hollows 2922 appears to create an ecological sink where breeding cannot occur (Isaac et al., 2014a). In 2923 our study of boobooks, nest site availability does not appear to be a limiting factor on 2924 abundance in urban and agricultural landscapes (Chapter 5). However, as demonstrated in 2925 Chapter 3, widespread and severe exposure to second generation anticoagulant 2926 rodenticides may reduce critical population parameters like survival and fecundity across 2927 age classes in areas of urban and periurban development. Determination of whether 2928 extensive conversion of bushland to urban or agricultural land uses creates an ecological 2929 sink or trap for boobooks would require quantification of population parameters including 2930 survival and fecundity in urban, agricultural, and bushland landscapes. However, 2931 degradation can likely be inferred from the dramatic differences in occupancy rates 2932 observed in Chapter 5.

2933 Care should be taken when interpreting the response of habitat generalists to novel 2934 anthropogenic land use types. Persistence in remnants surrounded by fragmentary matrix 2935 or even direct and consistent use of the matrix could indicate that the novel habitat type 2936 constitutes usable space of good or poor quality or it could indicate an ecological sink or 2937 trap. Distinguishing between these situations requires both knowledge of differences in 2938 demographic parameters between individuals using the fragmentary matrix and those using

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2939 intact natural habitat, as well as an understanding of patterns of habitat selection relative to 2940 the availability of the habitats on a scale approximating Johnson’s second order selection 2941 (Johnson, 1980). Future attempts to understand landscape-level impacts of key threatening 2942 processes should incorporate observations of survival and fecundity as well as the use of 2943 GPS telemetry or other techniques which enable the observation of individual habitat 2944 selection in order to facilitate quantification of these parameters.

2945 Additionally, attempts to assess impacts of habitat fragmentation on wildlife need to 2946 consider how individual species respond to different types of matrix. While habitat 2947 fragmentation clearly exerts pressure on some wildlife populations through direct habitat 2948 loss and small population phenomena impacting remaining isolated populations, 2949 threatening processes flowing from specific matrix types also need to be considered in 2950 modelling impacts of habitat alteration. Species with large home ranges may be especially 2951 vulnerable to unconventional edge effects, particularly when the threats involve pathogens 2952 and pollutants which can have substantial impacts on exposed individuals even when a 2953 relatively small portion of their home range includes the land use type where the threat 2954 originates (Lohr, 2018).

2955

2956 Management Recommendations

2957 Anticoagulant Rodenticides 2958 Species which are resilient to habitat fragmentation can sometimes compensate for 2959 habitat loss by utilizing resources in the surrounding matrix (Ewers and Didham, 2006). In 2960 the case of boobooks, which are generalist predators (Higgins, 1999), this resource subsidy 2961 may come largely in the form of high abundances of introduced bird species and commensal 2962 rodents in the urban matrix. Anticoagulant rodenticide poisoning, especially by more 2963 persistent SGARs poses a serious threat to boobooks with home ranges containing urban 2964 and suburban habitat (Lohr, 2018). Subsequent testing of a larger suite of carnivorous 2965 wildlife in Australia – including species listed under the Australian Commonwealth 2966 Environment Protection and Biodiversity Conservation Act 1999 – has revealed similar 2967 patterns across the continent (Michael Lohr, unpublished data). The pervasive use of SGARs 2968 in and around areas of human habitation threatens to convert potential matrix subsidies

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2969 into edge effects. Mitigating this threat will be important in maintaining boobooks and 2970 other carnivorous wildlife in urban systems and avoiding resultant tropic skew. In attempts 2971 to maintain native biodiversity in the face of extensive fragmentation the “loss of a few 2972 predator species often has impacts comparable in magnitude to those stemming from a 2973 large reduction in plant diversity” (Duffy, 2003). Bioaccumulation and biomagnification 2974 associated with highly-persistent SGARs make them particularly dangerous to non-target 2975 wildlife at the highest trophic levels and threaten to exacerbate trophic skew in already- 2976 susceptible urban ecosystems, hastening ecosystem decay.

2977 Regulatory restrictions have been implemented in the United States and Canada (Lohr 2978 and Davis, 2018). Despite implementation of restrictions in the United States in 2011 2979 (Bradbury, 2008), a recent study indicated that mean exposure in raptors had not declined 2980 (Murray, 2017). Removal of SGARs from sale directly to the public is probably necessary but 2981 not sufficient to prevent severe and widespread exposure in urban and exurban carnivores. 2982 I recommend complete replacement of currently used SGARs with commercially available 2983 less-persistent alternatives including baits based on the FGARs warfarin and coumatetralyl, 2984 cholecalciferol, or corn gluten meal. This regulatory reform should be coupled with 2985 increased research into effective alternative solutions to rodent control problems to ensure 2986 maintenance of a suite of effective rodent control options which reduce the probability of 2987 secondary poisoning on non-target wildlife.

2988 Nest Box Supplementation 2989 Nest boxes are a popular conservation intervention particularly among community 2990 groups and have been promoted for use to aid boobooks (Hussey, 1997). Nest boxes are 2991 intended to increase availability of nesting hollows where their abundance has been 2992 reduced by loss or alteration of native vegetation. My research suggests that nest hollow 2993 availability is not likely to be a limiting factor for boobooks in the urban and agricultural 2994 remnant bushlands where they were tested. Nest site availability may be more limiting in 2995 areas of intensive human land use where remnant bushlands are absent, but consideration 2996 needs to be given as to whether the addition of nest boxes may incentivise use of areas that 2997 are otherwise unsuitable, creating the potential for an ecological trap. A better 2998 understanding of population parameters and natural hollow availability in such areas is 2999 needed before advocating large-scale use of nest boxes as a conservation measure for

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3000 boobooks or other predatory bird species inhabiting areas of predominantly human land 3001 use.

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4092 Co-author Statements 4093

4094 Signed co-author statements verifying my role in the production of papers and manuscripts which 4095 make up chapters in this thesis are provided in this section.

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4104 Copies of original publications 4105 I include below copies of the first page of published peer-reviewed journal articles corresponding to 4106 chapters in this thesis. No licenses are required to reproduce these papers either in part or in full 4107 when included as part of a PhD thesis per the Elsevier license agreement: 4108 https://service.elsevier.com/app/answers/detail/a_id/565/track/AvMKOAoHDv8W~QaHGnwa~yKg_ 4109 38qZS75Mv9z~zj~PP_6/

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