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

ii

“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”

iii

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.

iv

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.

v

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

vi

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

vii

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.

viii

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.

ix

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).

x

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

xi

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

xii

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

xiii

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

xiv

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

xv

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

xvi

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

xvii

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.

1

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

2

(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

3

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

4

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

5

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.

6

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.

7

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

8

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.

9

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).

10

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.

11

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.

12

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 anti-

coagulant 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

13

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

14

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

15

(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).

16

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

Species Number Rodenticide Certainty State/Territory Source

Likely Exposure

Type Deitary

Category Reference

Reptiles King’s skink (Egernia

kingii) 8 brodifacoum physical symptoms Western Australia

island rat eradication primary omnivore Bettink, 2015

Birds Norfolk Island

Boobook (Ninox novaeseelandiae undulata) N/A brodifacoum suspected Norfolk Island

unspecified rat control program secondary carnivore Debus, 2012

Straw-necked Ibis (Threskiornis spinicollis) 1 bromadiolone

physical symptoms New South Wales

agricultural mouse control trial secondary

invertivore/carnivore Saunders, 1983

Barking Owl (Ninox connivens) 1 unknown

physical symptoms Queensland unknown secondary carnivore Thomas & Kutt, 1997

Barn Owl (Tyto alba) 1 brodifacoum

liver analysis (unknown concentration) Queensland

agricultural rat control secondary carnivore Thomas & Kutt, 1997

Lesser Sooty Owl (Tyto multipunctata) 2 brodifacoum

liver analysis (0.007 and <0.005 mg/kg) Queensland

agricultural rat control secondary carnivore Thomas & Kutt, 1997

Masked Owl (Tyto novaehollandiae) 1 brodifacoum

liver analysis (0.17 mg/kg) Queensland

agricultural rat control secondary carnivore Thomas & Kutt, 1997

Southern Boobook (Ninox novaeseelandiae) 1 unknown

museum record Queensland unknown secondary carnivore Thomas & Kutt, 1997

Brahminy Kite (Haliastur indus) 2 pindone suspected Western Australia

island rat eradication secondary carnivore Martin et al., 1994

17

Brown Falcon (Falco berigora) 1 unknown

physical symptoms Tasmania

private rodent control secondary carnivore Mooney, 2017

Brown Goshawk (Accipiter fasciatus) 2 unknown

physical symptoms Tasmania

private rodent control secondary carnivore Mooney, 2017

Collared Sparrowhawk (Accipiter cirrocephalus) 1 unknown

physical symptoms Tasmania

private rodent control secondary carnivore Mooney, 2017

Grey Goshawk (Accipiter novaehollandiae) 5 unknown

physical symptoms Tasmania

private rodent control secondary carnivore Mooney, 2017

Tasmanian Masked Owl (Tyto novaehollandiae castanops) 12 unknown

physical symptoms Tasmania

private rodent control secondary carnivore Mooney, 2017

Tasmanian Boobook (Ninox novaeseelandiae leucopsis) 6 unknown

physical symptoms Tasmania

private rodent 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 philippensis) 5 brodifacoum

physical symptoms Western Australia

island rat eradication primary invertivore Palmer, 2014

Silver Gull (Larus novaehollandiae) 7 brodifacoum

physical symptoms Western Australia

island rat eradication both

invertivore/carnivore Palmer, 2014

18

Pacific Golden Plover (Pluvialis fulva) 1 brodifacoum suspected Western Australia

island rabbit eradication both invertivore Palmer, 2014

Ruddy Turnstone (Arenaria interpres) 28 brodifacoum

physical symptoms Western Australia

island rat eradication secondary invertivore Palmer, 2014

Buff-banded Rail (Gallirallus philippensis) 2 brodifacoum suspected New South Wales

island rabbit eradication

not specified omnivore Priddel et al., 2000

Pied Currawong (Strepera graculina) 1 brodifacoum suspected New South Wales

island rabbit eradication

not specified omnivore Priddel et al., 2000

Little Raven (Corvus mellori) 1 bromadiolone

physical symptoms Victoria

residential rodent control

not specified omnivore Reece et al., 1985

Purple Swamphen (Porphyrio porphyrio melanotus) 1 bromadiolone

physical symptoms Victoria

residential rodent control

not specified omnivore Reece et al., 1985

Brown Skua (Stercorarius antarcticus lonnbergi) 512 brodifacoum

physical symptoms Tasmania

island rabbit eradication secondary carnivore

Tasmania Parks and Wildlife Service, 2014

Kelp Gull (Larus dominicus) 988 brodifacoum

physical symptoms Tasmania

island rabbit eradication primary

invertivore/carnivore

Tasmania Parks and Wildlife Service, 2014

Northern Giant Petrel (Macronectes giganteus) 693 brodifacoum

physical symptoms Tasmania

island rabbit eradication secondary carnivore

Tasmania Parks and Wildlife Service, 2014

Pacific Black Duck (Anas superciliosa superciliosa) and Mallard (A. platyrhynchos platyrhynchos) 157 brodifacoum

physical symptoms Tasmania

island rabbit eradication primary omnivore

Tasmania Parks and Wildlife Service, 2014

Southern Giant Petrel (Macronectes halli) 38 brodifacoum

physical symptoms Tasmania

island rabbit eradication secondary carnivore

Tasmania Parks and Wildlife Service, 2014

Unknown Bird 5 brodifacoum physical Tasmania island rabbit not

Tasmania Parks and

19

symptoms eradication specified Wildlife Service, 2014

Unknown giant petrel (Macronectes sp.) 31 brodifacoum

physical symptoms Tasmania

island rabbit eradication secondary carnivore

Tasmania Parks and 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 longimembris) 1 brodifacoum liver analysis Queensland

agricultural rat control secondary carnivore Young & De Lai, 1997

Masked Owl (Tyto novaehollandiae) 1 brodifacoum

physical symptoms Queensland

agricultural rat control secondary carnivore Young & De Lai, 1997

Rufous Owl (Ninox rufa) 2 brodifacoum

physical symptoms Queensland

agricultural rat 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 vulpecula) 7 unknown

physical symptoms Queensland unknown

not specified omnivore Grillo et al., 2016

boodie (Bettongia lesueur) 20-50 pindone

population eradicated Western Australia

island rat eradication primary herbivore Morris, 2002

20

In addition to accounts of wildlife poisoning, we also located published accounts 2

suggesting population-level effects of rodenticide toxicity on carnivorous birds in Australia. 3

Olsen (1996) listed the use of rodenticides in areas of palm cultivation as a potential 4

contributing factor in the decline of Norfolk Island Boobooks (Ninox novaeseelandiae 5

undulata x novaeseelandiae). Young and De Lai (1997) observed a correlation between 6

declines in owl abundance and the use of “Klerat®”a brodifacoum-based rodenticide in 7

sugar cane fields in north Queensland and documented one confirmed and several 8

suspected cases of brodifacoum poisoning in owls (James, 1997). A subsequent report 9

noted three additional cases of owls in Queensland testing positive for brodifacoum 10

residues (0.007 mg/kg, <0.005mg/kg, and 0.17mg/kg ) in the 1990s and two museum 11

specimens of Southern Boobooks (Ninox novaeseelandiae) with rodenticide poisoning listed 12

as their cause of death in the collection notes (Thomas and Kutt, 1997). One of the two 13

specimens, while alive showed symptoms of AR poisoning including “bleeding from the 14

nasal passages; loss of muscle co-ordination; lethargy including drooping head and eyes; 15

and generally poor and dirty condition” (Thomas and Kutt, 1997). The report reviewed 16

several other factors which could potentially have impacted owl populations in the area and 17

came to the conclusion that there was “significant potential for secondary poisoning of owls 18

to occur in Queensland sugarcane as a result of the use of Klerat®” (Thomas and Kutt, 1997). 19

Crop Care Australia later deregistered Klerat® for use in sugar cane fields over concerns 20

relating to secondary poisoning (Twigg et al., 1999). 21

An unpublished PhD dissertation examined dynamics of secondary poisoning of 22

avian predators associated with sugar cane fields in Queensland and concluded that the 23

coumatetralyl-based product used to control rats did not pose a threat to predatory birds 24

(Ward, 2008). This conclusion was based largely on the low relative use of canefields for 25

foraging by predatory birds, the low concentration of coumatetralyl in rats captured outside 26

of canefields, and the low toxicity and persistence of coumatetralyl relative to second 27

generation anticoagulant rodenticides (Ward, 2008). Unfortunately, no predatory birds in 28

the treated areas were directly tested for rodenticide exposure. A lack of detection of 29

coumatetralyl in Southern Boobooks in Western Australia as part of an ongoing study 30

supports the low probability of secondary toxicity in raptors for this rodenticide. 31

21

Pindone has been implicated as a factor driving the decline of Little Eagle (Hieraaetus 32

morphnoides) numbers in and around Canberra (Olsen et al. 2013). Breeding pairs of Little 33

Eagles disappeared from areas baited with pindone while pairs in areas baited with 1080 or 34

not baited at all persisted (Olsen et al. 2013). The high susceptibility of Wedge-tailed Eagles 35

to pindone in laboratory tests (Martin et al. 1994) lends credibility to the hypothesis that 36

pindone could be responsible. Unfortunately, no direct testing of Little Eagles suspected of 37

being poisoned was conducted to confirm pindone exposure and rule out other ARs from 38

residential and commercial sources. 39

Recently, a study in Tasmania examined probable rodenticide poisoning in predatory 40

birds. Six species (Table 2.2) showed signs of anticoagulant rodenticide poisoning when 41

dissected (Mooney, 2017) but the rodenticides responsible were not determined or 42

quantified. As part of this study, thirteen predatory bird species were ranked by risk of 43

rodenticide exposure according to four natural history parameters: relative metabolic 44

speed, dietary habits influencing consumption of contaminated tissues, relative preference 45

for rodents, and willingness to forage near anthropogenic structures (Mooney, 2017). 46

Development of a more statistically robust predictive model using similar natural history 47

parameters to examine risk of rodenticide exposure in a wider range of predatory species 48

would be an extremely useful step toward assessing likely population level impacts on 49

wildlife in Australia. Incorporating variables relating to seasonal dietary shifts and home 50

range size could potentially improve future models. 51

The overall lack of attention within Australia to what is perceived as a potentially 52

serious threatening process for native carnivores in many other parts of the world suggests 53

the need for Australian studies which examine potential impacts on native fauna in a 54

quantitative and comprehensive manner. Susceptibility of marsupial carnivores is 55

particularly poorly understood and should be a focus of future research. Furthermore, a 56

surveillance program should be in place in areas of high AR use, to monitor any dead wildlife 57

for a cause of death. Most of the studies we used in this review did not sample animals and 58

thus were not able to confirm suspicions of death due to rodenticide poisoning. 59

4.3 Governance and legislation of rodenticide use 60

22

At present, no information is available on the volume of sales or application of ARs in 61

Australia. Reporting for all poisons intended to control vertebrates indicates that 222 62

different products are currently registered with a total sales reaching $18,601,875.00 in the 63

2015-2016 fiscal year (Australian Pesticides and Veterinary Medicines Authority, 2017a). 64

Nine anticoagulants are currently approved for vertebrate pest control in Australia (McLeod 65

and Saunders 2013). At present, all nine are listed as Schedule 6 substances (see Appendix 66

2.A for schedule meanings) in Australia (Australian Government Department of Health: 67

Therapeutic Goods Administration, 2017) (Table 2.3) and are legally allowed to be sold 68

directly to the public and do not require government permits for purchase or use. In some 69

cases, more concentrated formulations of SGARs are listed as Schedule 7 substances and are 70

restricted to licensed pesticide applicators (Australian Government Department of Health: 71

Therapeutic Goods Administration, 2017) while products containing low concentrations of 72

some FGARs are registered as schedule 5 substances which require only simple warnings 73

and safety directions for public sale (Table 2.3). The FGAR diphacinone is currently approved 74

as an active ingredient but has no products registered with the APVMA after July 2016 75

(Australian Pesticides and Veterinary Medicines Authority, 2017b). However, remaining 76

stock can still be used for 12 months following a stopped registration (Commonwealth of 77

Australia, 1994) and MSDS sheets obtained from a pest management contractor seem to 78

indicate that at least one diphacinone product is still in use at present. The APVMA has 79

prioritised a review of the status of all SGARs currently approved in Australia (brodifacoum, 80

bromadiolone, difenacoum, difethialone, and flocoumafen) citing concerns over public 81

health, worker safety, and environmental safety (Australian Pesticides and Veterinary 82

Medicines Authority, 2015)83

23

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

Anticoagulant Chemical Class Generation Schedule (See Appendix 2.A)

Acute Oral LD50 (Rattus norvegicus)

mg/kg

LD50 Reference Approved

Target Species

brodifacoum hydroxycoumarins second 6 (0.25 per cent or less) or 7 0.27

Godfrey 1985 mice and rats

bromadiolone hydroxycoumarins second 6 (0.25 per cent or less)or 7

0.57-0.75 Meehan 1978 mice and rats

coumatetralyl hydroxycoumarins first 5 ( 0.05 per cent or less), 6 (1 per cent or less), or 7 16.5

Dubock and Kaukeinen 1978 mice and rats

difenacoum hydroxycoumarins second 6 (0.25 per cent or less) or 7 1.8-3.5

Bull 1976 mice and rats

difethialone hydroxyl-4-benzothiopyranones

second 6 (0.0025 per cent or less) or 7 0.27-0.69

Lechevin and Poche 1988 mice and rats

diphacinone indandiones first* 6 1.93-2.7 Fisher et al. 2003 approval expired

flocoumafen hydroxycoumarins second 6 (0.005 per cent or less) or 7

0.25-0.56 Lund 1988 mice and rats

pindone indandiones first* 6 75-100 Fisher et al. 2003 rabbits

warfarin hydroxycoumarins first 4†, 5 (0.1 per cent or less), or 6 3.3

Fisher et al. 2003 mice and rats

24

Increasing concerns over risks to the health and safety of humans and pets and 86

impacts on non-target wildlife have prompted stricter regulation of anticoagulant 87

rodenticides – particularly SGARs – in several developed nations. While rodenticide 88

legislation is often complex and varies substantially between countries, the trend is toward 89

stricter legislation than currently exists in Australia. In the United States, SGARs are 90

restricted to licensed pesticide applicators, only allowed to be used indoors, and are 91

required to be placed in containers which exclude children and pets (Bradbury, 2008). 92

Similar requirements were subsequently implemented in Canada (Health Canada: Pest 93

Management Regulatory Agency, 2010). A somewhat different approach is taken in the UK, 94

where SGARS are licensed for outdoor use but an industry taskforce has been established to 95

monitor both rodenticide applicator usage patterns and breeding success and SGAR residues 96

in the livers of one sentinel species – Barn Owls (Tyto alba) – to determine the impacts of 97

this legislative change on exposure rates (Shore et al., 2016). These alternative models of 98

AR regulation and the direction they represent in evolving global norms should be 99

considered when evaluating current Australian regulations. 100

Given the changes in legislation governing the use of ARs in other developed nations 101

and demonstrated impacts on human health and wildlife populations overseas, we support 102

the ongoing review of the use and scheduling of SGARs in Australia by the APVMA. In 103

Australia, AR poisoning has been documented in pets (Robertson et al., 1992) and humans 104

(Osborne et al., 2017), particularly children (Ozanne-Smith et al., 2001; Parsons et al., 1996; 105

Reith et al., 2001). Roughly 1,400 human exposures to ARs per year are recorded by Poison 106

Information Centres in Australia (Australian Pesticides and Veterinary Medicines Authority, 107

2015). Removal of SGARs from retail sale to the public by listing all SGARs as schedule 7 108

poisons and implementing stricter requirements that baits be used only indoors and placed 109

in a manner that makes them inaccessible to children and pets will help to bring Australian 110

practices closer to emerging global norms and best practices. These actions are likely to 111

help to mitigate human health and safety risks and exposure in non-target wildlife. Critical 112

evaluation of whether these practices are effective will require long-term monitoring of AR 113

residues in appropriate sentinel species – as practiced in the UK – before and after any 114

regulatory changes are implemented. Ongoing research into exposure patterns in Southern 115

25

Boobooks will provide valuable baseline data for a widely-distributed sentinel species if the 116

suggested regulatory changes are implemented. 117

Current Uses in Australia 118

Agricultural 119

In Australian agriculture, ARs are primarily used in asset protection around 120

infrastructure and grain storage areas and many first and second generation products are 121

licensed for these purposes. In the past, several trials have been conducted on broadscale 122

application of rodenticides in Australian cropping systems. 123

Brown & Singleton (1998) found aerial distribution of brodifacoum-based baits 124

effective at controlling mice in wheat fields in South Australia in a field trial and the authors 125

suggested that application according to guidelines was unlikely to cause substantial non-126

target mortality. However, mice were observed to be active during the day following the 127

baiting, which the authors acknowledged could increase the risk of secondary poisoning in 128

predatory species (Brown and Singleton, 1998). To our knowledge, aerial distribution of 129

brodifacoum baits in wheat crops has never been implemented on an operational basis in 130

Australian agriculture. 131

Several trials of bromadiolone efficacy in controlling mouse plagues have been 132

conducted in agricultural crops in Australia. In the earliest of these studies, aerial 133

application was used to distribute bromadiolone bait directly into sunflower crops in New 134

South Wales (Saunders, 1983). Bromadiolone was identified as the most promising of the 135

three toxicants tested but the authors noted concern over bromadiolone’s slow method of 136

action potentially facilitating secondary poisoning of predators selecting for poisoned mice 137

(Saunders, 1983). One Straw-necked Ibis (Threskiornis spinicollis) was found dead of 138

apparent rodenticide poisoning after having consumed 6-10 mice in an area where 139

bromadiolone had been aerially applied as part of a trial to control mice in sunflower crops 140

(Saunders, 1983). In a subsequent study, wheat laced with bromodialone was applied a 141

single time in bait stations in soybean crops in New South Wales (Twigg et al., 1991). The 142

study did not search for or detect any mortalities in non-target wildlife but cautioned that 143

“The risks to non-target species and of contaminating primary produce posed by broad-scale 144

use of rodenticides would need to be assessed fully before these chemicals could become 145

26

an integral part of farm management. In Australia, such data are sparse and research is 146

required urgently” (Twigg et al., 1991). The only subsequent available study on broad-scale 147

use of bromadiolone in agriculture used a fertiliser spreader to apply four treatments of 148

wheat laced with bromadiolone to “refuge habitat, channel banks, fence lines, non-arable 149

land and road verges” within 200m of soybean crops in New South Wales but failed to 150

demonstrate significant reductions in crop damage (Kay et al., 1994). It does not appear 151

that non-target exposure was evaluated as part of this study. 152

Contrary to the warning issued by Twigg et al. (1991), which cautioned a more 153

complete assessment of non-target impact prior to the broad-scale use of ARs in agriculture, 154

under some circumstances, ARs are or have been used in or adjacent to crops to control 155

mice and rats. During mouse plagues, temporary registrations for the use of bromadiolone 156

have been issued for use in wheat crops in Victoria in 1984, perimeter baiting of oilseed 157

crops in New South Wales in 1984-1985, and in soybean crops in New South Wales in 1989 158

(Twigg et al., 1991). Expired permits issued to allow the baiting of crop perimeters with 159

bromadiolone show valid periods between 16 September 1999 and 31 December 1999 160

(PER3031); 06 December 2006 and 30 March 2009 (PER9543); and 31 March 2009 and 30 161

June 2016 (PER11331) (Australian Pesticides and Veterinary Medicines Authority, 2017b). 162

There are no current permits for the use of bromadiolone in perimeter baiting around crops 163

but a current New South Wales government factsheet and web page state that 164

bromadiolone bait can be prepared by the Livestock Health and Pest Authority (LHPA) for 165

availability to farmers in perimeter baiting around crops (New South Wales Department of 166

Primary Industries, 2011; New South Wales Government: Department of Primary Industries, 167

2017). 168

The SGAR brodifacoum was also previously applied broadscale in sugar cane fields in 169

Queensland (Young and De Lai, 1997) but the registration for that use has since been 170

revoked over concerns about mortality in non-target wildlife (Twigg et al., 1999). The use of 171

brodifacoum in this context has largely been replaced by the use of the FGAR coumatetralyl. 172

Research on non-target impacts of coumatetralyl in sugar cane fields demonstrated low risk 173

of secondary toxicity (Ward, 2008). Coumatetralyl is currently registered for use in 174

pineapple, macadamia, and sugar cane crops in all states and territories (Australian 175

Pesticides and Veterinary Medicines Authority, 2017b). 176

27

Published literature and official accounts may seriously underestimate the usage of 177

ARs in cropping systems in Australia. A study of second generation anticoagulant use in 178

agricultural systems in Northern Ireland found that total compliance with best practice 179

application methods was rare and lack of compliance probably facilitated greater risk of 180

secondary toxicity to native wildlife (Tosh et al. 2011). Within Australia, landowners have 181

requested that a government agency provide them with pindone with the intention of using 182

pindone to reduce kangaroo abundance in contravention of its label (Twigg et al., 1999). 183

Many ARs are readily available in hardware and agricultural supply stores in Australia 184

without a permit and the potential for use contrary to labelling restrictions is high. A better 185

understanding of current legal and illegal usage of ARs in agriculture is necessary to 186

determine the likelihood of secondary poisoning of non-target species in agricultural 187

systems. 188

Conservation 189

ARs have a long history of use on islands and in fenced reserves worldwide for 190

eradication of rodents for conservation purposes. At present, application of ARs is the only 191

effective way of removing introduced rodents from islands larger than 5ha for conservation 192

purposes (Campbell et al., 2015). Many successful and well-documented eradications of 193

introduced rodents and rabbits have been conducted in Australia using ARs (Bettink, 2015; 194

Burbidge, 2004; Cory et al., 2011; Dunlop et al., 2015; Meek et al., 2011; Morris, 2002; 195

Priddel et al., 2000; Tasmania Parks and Wildlife Sevice, 2014). Pindone was used in some 196

early eradications but its use has largely been supplanted by brodifacoum (Burbidge and 197

Morris, 2002) and bromadiolone (Meek et al., 2011). Reviews of island eradications have 198

been conducted for New South Wales (Priddel et al., 2011) and Western Australia (Burbidge 199

and Morris, 2002). 200

During the course of some eradications, high levels of non-target mortality and 201

poisoning of species listed under the Australian Environment Protection and Biodiversity 202

Conservation Act 1999 have been documented. In one instance, boodies (Bettongia lesueur) 203

(listed as vulnerable) were accidentally eradicated on Boodie Island along with the intended 204

target, black rats (Rattus rattus) (Morris, 2002). An eradication of black rats was proposed 205

for Woody Island in Western Australia but was halted when the rats on the island were 206

subsequently identified as a native species (Rattus fuscipes) (Burbidge et al., 2012). During 207

28

the successful eradication of rabbits, black rats, and mice (Mus musculus) on Macquarie 208

Island, concerns were expressed by the public and government authorities over the 209

observed mortality of 2,424 individuals from several seabird and waterfowl species, 210

presumably related to the use of the SGAR brodifacoum (Tasmania Parks and Wildlife 211

Sevice, 2014). While some species, especially Northern Giant Petrels (listed as vulnerable) 212

experienced substantial population-level declines as a result of the baiting, the reductions 213

were expected to be temporary and removal of introduced mammals has already facilitated 214

improved population parameters in a number of seabird species (Tasmania Parks and 215

Wildlife Sevice, 2014). Endangered Southern Giant Petrels were also lethally poisoned 216

during the course of this eradication (Tasmania Parks and Wildlife Sevice, 2014). Collateral 217

damage to non-target species may be acceptable and necessary in some situations but more 218

careful consideration and planning are required to avoid poor outcomes which have 219

occurred or been narrowly averted during rodent eradications in the past. In some 220

instances, bait boxes modified to exclude native fauna may decrease the incidence of 221

primary of non-target wildlife AR exposure during eradication attempts (Moro, 2001). Use 222

of biological control agents prior to baiting can also increase the probability of success and 223

reduce the volume of poison needed to remove target animals (Priddel et al., 2000). Close 224

monitoring of non-target mortality during and after island eradications is necessary to 225

properly assess the relative benefit to native biodiversity. 226

In Australia, ARs have also been tested as a method to control feral pigs for 227

conservation purposes and reduction of agricultural threats. Trials using the FGAR warfarin 228

were conducted in New South Wales (Choquenot et al., 1990; Saunders et al., 1990) and the 229

Australian Capital Territory (McIlroy et al., 1989). While two of the three trials found the 230

use of warfarin to be highly effective, this method does not appear to have been put into 231

practice due to concerns over animal ethics, non-target exposure, and a shift toward the use 232

of 1080 baits for pig control (Cowled et al., 2008). However, the use of warfarin to control 233

feral pigs in Australia has been recommended in the published literature as recently as 2014 234

(McIlroy, 2014). While warfarin is unlikely to cause secondary poisoning in exposed wildlife, 235

the risk of primary poisoning to wildlife consuming bait intended for pigs is likely too high to 236

warrant the use of this method of control. 237

29

Residential and Commercial 238

Patterns of residential and commercial use of ARs in Australia are poorly known. At 239

present, the Australian Pesticide and Veterinary Medicine Association (APVMA) lists seven 240

ARs (two FGARs and five SGARs) as registered for use in Australia in commercial and 241

residential settings (Table 2.3). We have observed two FGARs (warfarin and coumatetralyl) 242

and three SGARs (brodifacoum, bromadiolone, and difenacoum) available for purchase by 243

the public at retail outlets in Western Australia. The SGARs flocoumafen and difethialone 244

are also used by commercial pest control companies in residential and commercial settings. 245

Residues of both have been detected in native wildlife in Western Australia. Patterns of 246

availability to unlicensed individuals are similar to those in the UK where three FGARs and 247

five SGARs are registered for use and are not restricted to licensed applicators (Shore et al., 248

2016). However, regulations governing AR use are substantially more restrictive in some 249

other industrialized countries. In the US, three FGARs are permitted for use by the public 250

but all four registered SGARs are restricted to use by licensed pesticide applicators 251

(Bradbury, 2008). Similarly, in Canada the public has access to three FGARs and licensed 252

contractors may use an additional three SGARs (Health Canada: Pest Management 253

Regulatory Agency, 2010). 254

The lack of available data on the quantities of ARs used in domestic and commercial 255

settings and the locations where they are used makes it nearly impossible to gauge the 256

potential non-target impacts of these products. Only two publications directly implicate 257

private use of rodenticides in non-target mortality in Australia. In the most definitive 258

example, brodifacoum was implicated in the deaths of a Purple Swamphen (Porphyrio 259

porphyrio melanotus) and Little Raven (Corvus mellori) which showed signs of AR poisoning 260

after baiting in a residential area (Reece et al., 1985). In Tasmania, residential and small-261

scale agricultural baiting is thought to have been the source of ARs responsible for the 262

suspected lethal poisonings of 27 individuals from six raptor species (Mooney, 2017). Given 263

that use of rodenticides in conservation and agricultural contexts is relatively limited and 264

only occurs periodically, the total amount deployed in residential and commercial settings is 265

likely to be far greater. Accordingly, overseas studies on rodenticide exposure in bobcats 266

(Lynx rufus) in America (Riley et al., 2007) and a variety of bird and mammal species in Spain 267

(López-Perea et al., 2015) indicate a spatial correlation between population density and AR 268

30

exposure in wildlife. Collection of basic information on the quantities of ARs sold to private 269

residents and pest control contractors by locality coupled with systematic testing of wildlife 270

populations across different land-use types will be essential in assessing the risks posed to 271

non-target wildlife by residential and commercial use of ARs. 272

Unique Considerations in Australia 273

Pindone 274

Unlike other ARs used in Australia, the SGAR pindone has received more scrutiny and 275

has been the focus of a greater body of research because of its longer history of use and 276

large scale of use in rabbit control. At present, it is only registered for use in Australia and 277

New Zealand (P. Fisher, Brown, & Arrow, 2015; Twigg et al., 1999) and, as a consequence, 278

has received little attention by researchers elsewhere in the world. Efficacy trials for rabbit 279

control were conducted in Western Australia in 1971-1975 (Oliver et al., 1982) and 1981-280

1982 (Robinson and Wheeler, 1983). Pindone was registered in Western Australia for rabbit 281

control in 1984 and was subsequently registered for the same use in all other Australian 282

states (Twigg et al., 1999). Pindone was registered for use in New Zealand in 1992 (Twigg et 283

al., 1999). In Australia, pindone is used in rabbit control primarily in areas where the use of 284

sodium fluoroacetate (1080) is deemed to pose too great a risk to humans and pets 285

(Department of Agriculture and Food Western Australia, 2015). Such areas include “market 286

gardens, golf courses, hobby farms, around farm buildings” (Twigg et al., 1999) and 287

bushlands adjacent to populated areas. In the past, it has also been used in island 288

eradications of rabbits and rodents prior to being largely replaced by brodifacoum (Burbidge 289

and Morris, 2002; Priddel et al., 2011). 290

Pindone use in Australia has been the subject of extensive review (National 291

Registration Authority For Agricultural and Veterinary Chemicals, 2002; Twigg et al., 1999) 292

prompted by public concern over reports of lethal poisoning of non-target species (Table 293

2.2). As a consequence, additional restrictions were placed on the sale of pindone 294

concentrates and labelling was required to include a “statement not to lay baits in the 295

vicinity of native animal habitat” (National Registration Authority For Agricultural and 296

Veterinary Chemicals, 2002). 297

31

At present, little is known about the effects of pindone on non-target species. 298

Pindone has been shown in laboratory tests to have varying effects on different native 299

Australian bird taxa (Martin et al. 1994). Wedge-tailed Eagles were more susceptible than 300

other species tested but Common Bronzewings (Phaps chalcoptera) and other granivores 301

were also noted to be at high risk of poisoning due to direct consumption of poisoned grain 302

(Martin et al. 1994). Despite the authors’ recommendation for field studies of impacts on 303

Wedge-tailed Eagles and other raptors (Martin et al., 1994), to the best of our knowledge, 304

no further study on this topic has been conducted in Australia. 305

The repeated use of an anticoagulant in natural areas to control but not eradicate 306

rabbits appears to be unique to Australia and New Zealand. The repeated pattern of use in 307

the same areas may pose a serious long-term threat to susceptible wildlife populations. This 308

may be especially problematic for long-lived species with low reproductive rates which are 309

unable to sustain low levels of additive mortality. The potential link between pindone 310

baiting and the decline of Little Eagles in Canberra (Olsen et al., 2013) exemplifies this 311

concern. However, an ongoing study of rodenticide exposure in Southern Boobooks has not 312

detected any pindone residue in samples tested to date despite testing of samples obtained 313

in areas where pindone baiting has occurred. Differences in diet, territory size, and 314

metabolism could account for this lack of detection. In some instances, reduction of prey 315

abundance via ARs could potentially drive declines in predatory species rather than direct 316

ARs toxicity. However, in the instance of Little Eagles in Canberra, this does not appear to 317

be the case, as the decline of Little Eagle abundance was independent of rabbit abundance 318

(Olsen et al., 2013). Additional research into the sensitivity of Australian fauna to pindone 319

and the population impacts of different patterns of use are necessary to determine the 320

extent and severity of impacts on non-target fauna. At minimum, the continued use of 321

pindone to control rabbits in bushland areas needs to be evaluated as to whether it 322

provides a net benefit or detriment to the conservation of native biodiversity. 323

Human consumption of rabbits is common in agricultural areas and may facilitate 324

some risk of human exposure to pindone. Risk of substantial human exposure is reduced by 325

the fact that livers are not typically consumed. However, pindone has been demonstrated 326

to accumulate in fat tissue in rabbits at similar concentrations to liver tissue (Fisher et al., 327

2015). Some discussions of risk of human exposure to ARs via ingestion of contaminated 328

32

game meats have suggested that cooking prior to consumption might reduce AR exposure 329

through degradation of the relevant chemicals (Eisemann and Swift, 2006). Conversely, 330

subsequent empirical research demonstrated that, at least in pig tissues contaminated with 331

diphacinone, cooking did not substantially reduce AR concentration (Pitt et al., 2011). While 332

we consider the risk of pindone poisoning associated with human consumption of wild 333

rabbits to be low due to its relatively short half-life and low acute toxicity, as a minimum 334

precaution we recommend adhering to established 5 week withholding period for livestock 335

exposed to pindone (Twigg et al., 1999). 336

Reptiles 337

We found only one example of documented or suspected lethal AR poisoning of 338

reptiles in Australia (Bettink, 2015) in the course of our literature search. A further 339

investigation of international literature revealed serious gaps in knowledge relating to 340

impacts of ARs on reptiles and their potential role as vectors to higher trophic levels. In 341

combination, the few existing published accounts suggest that some reptiles may be more 342

resistant to anticoagulant rodenticides than birds or mammals. As a consequence, 343

developing a better understanding of how reptiles are impacted by AR exposure and their 344

potential as vectors to more vulnerable taxa will be critical to evaluating the ecotoxicology 345

of ARs in areas of the world where reptiles are a substantial component of biodiversity. 346

The mechanisms by which carnivorous birds and mammals are exposed to ARs have 347

not been widely researched (Elliott et al., 2014). The few studies investigating AR exposure 348

in intermediate vectors tend to focus on insects (Masuda et al., 2014), and small mammals 349

(Brakes and Smith, 2005) as potential vectors (Elliott et al., 2014) with the vast majority of 350

work focusing on target and non-target small mammals (Hoare and Hare, 2006). Because 351

most of these studies have been conducted in temperate areas of Europe or North America, 352

they may not be representative of dominant exposure pathways in tropical and warm arid 353

areas of the world. In areas where reptiles are more diverse and abundant, reptiles may act 354

as an important pathway for transmission of ARs through terrestrial food webs because of 355

their increased relative importance as prey items for carnivores at higher trophic levels 356

(Hoare and Hare, 2006). Furthermore, in ecosystems with a high predominance of 357

carnivorous reptiles e.g. snakes, monitor lizards and large skinks, there may be a direct bio-358

33

accumulation effect when reptiles prey on rats or mice directly, or on other reptiles, leading 359

to a negative impact on larger-bodied reptiles (Bishop et al., 2016; Olsson et al., 2005). 360

Reptiles make up a substantial proportion of the prey base of some carnivores in 361

Australia (Doherty et al., 2015; Paltridge, 2002) and comprise >80% of the biomass in the 362

diets of some predatory bird species (Aumann, 2001). Reptile diversity and abundance is 363

substantially higher in Australia than in Europe and North America (Roll et al., 2017) where 364

secondary anticoagulant rodenticide exposure has been more comprehensively assessed in 365

native fauna. As a consequence, understanding patterns of exposure in reptiles and their 366

capacity to transmit ARs to higher trophic levels is critical to understanding ecosystem level 367

AR exposure in Australia and other countries with high reptile abundance. Only a few 368

studies have investigated the mechanisms and ramifications of AR exposure in reptiles 369

(Hoare and Hare, 2006). In one instance, the SGAR brodifacoum was detected in Pinzón lava 370

lizards (Microlophus duncanensis) up to 850 days after baiting of an uninhabited island with 371

no other rodenticide sources (Rueda et al., 2016). Long duration of AR persistence in lava 372

lizards could be a consequence of recursive exposure from consumption of invertebrates 373

feeding on reptile faeces containing AR residue, low elimination rates by lizards, or slow 374

decomposition leading to prolonged availability of bait (Rueda et al., 2016). Subsequent 375

deaths of 22 Galapagos hawks (Buteo galapagoensis) showing signs of rodenticide toxicity 376

were attributed to secondary poisoning resulting from consumption of lava lizards, as was 377

the death of a short-eared owl (Asio flammeus) found dead with lethal concentrations of 378

brodifacoum present in its liver 773 days after baiting (Rueda et al., 2016). If other reptile 379

species are also capable of vectoring lethal levels of rodenticide to higher trophic levels for 380

greater than two years after initial exposure, the threat of secondary poisoning to 381

carnivorous birds and mammals in regions of the world with diverse and abundant 382

herpetofaunas may be severely underestimated. 383

High tolerance to AR exposure may also increase the efficacy of reptiles as vectors of 384

ARs to higher trophic levels. At least some reptiles appear to be substantially more resistant 385

to AR toxicity than birds or mammals (Weir et al., 2015). An acute oral LD50 of 550 μg/g 386

was determined for the AR pindone in Western fence lizards (Sceloporus occidentalis) (Weir 387

et al., 2015). No LD50 was determined for the SGAR brodifacoum because all western fence 388

lizards tested survived the highest does of 1,750 μg/g (Weir et al., 2015). Both LD50s are 389

34

three to five orders of magnitude higher than in most bird and mammals species tested 390

(Laakso et al., 2010). Similarly, when prairie rattlesnakes (Crotalus viridis) were fed three 391

laboratory mice poisoned with bromadiolone over the course of three weeks, none of the 392

snakes died or showed signs of rodenticide toxicity in the 30 days following the treatment 393

despite consuming more mg/Kg brodifacoum than the LD50s established for several 394

mammal species in the same study (Poché, 1988). Pitt et al. (2015) examined brodifacoum 395

residues in 112 geckoes (Lepidodactylus lugubris and Hemidactylus frenatus) collected on 396

Palmyra Atoll after rat control operations. They noted a peak concentration of 0.067 μg/g 397

and detectable concentrations at about half of this rate were still noted 60 days post-baiting 398

Pitt et al., 2015). Pitt et al. (2015) concluded that geckos were unlikely to experience 399

mortality but on islands where secondary predators existed, there could be some 400

ecosystem-wide impacts. Similarly, bungarras or Gould’s goannas (Varanus gouldii) were 401

observed consuming rats poisoned with brodifacoum during an eradication in the 402

Montebello Islands of Western Australia, but did not appear to experience adverse effects 403

(Burbidge, 2004). If a tolerance for rodenticides exists across multiple reptile taxa, reptiles 404

may be more effective at concentrating and transmitting ARs to higher trophic levels than 405

the small mammals which have been more commonly examined as potential vectors of ARs 406

to higher trophic levels. 407

Conversely, in some instances, apparent susceptibility of some reptile species to ARs 408

has been observed or hypothesized. In Australia, the single documented account of lethal 409

AR toxicity in reptiles involved the direct ingestion of brodifacoum baits by King’s skinks 410

(Egernia kingii) during a rat eradication on Penguin Island in Western Australia (Bettink, 411

2015). Eight of the skinks were found dead and exhibited haemorrhage associated with AR 412

toxicity and several others were treated with vitamin K and released (Bettink, 2015). 413

Subsequent analysis revealed a concentration of 1.3 mg/kg in the liver of one of the dead 414

skinks (Bettink, 2015). This liver concentration is well above minimum lethal thresholds 415

suggested for many bird and mammal species so it is difficult to infer relative susceptibility 416

of King’s skinks from this event. Sánchez-Barbudo et al. (2012) documented the death of a 417

horseshoe whip snake (Hemmorrhois hippocrepis) due to flocoumafen used to protect a 418

seabird colony. A number of anecdotal accounts of lethal AR poisoning have also been 419

reported in skinks and geckos (Wedding et al., 2010). Susceptibility of goannas in Australia 420

35

to poisoning with brodifacoum has also been suggested (James, 1997), although it appears 421

that this only considers the likelihood of exposure due to carrion being a component of their 422

diet rather than an actual vulnerability to the effects of brodifacoum. The lack of observed 423

mortality in some reptile species may be due to a delayed onset of effects relative to birds 424

and mammals. This possibility is supported by the observation of the deaths of six 425

Galápagos land iguanas (Conolophus subcristatus) more than two months after their island 426

was baited with brodifacoum to control rats. Merton (1987) described a similar incident in 427

which Telfair’s skinks (Leiolopisma telfairii) were found dead three to six weeks after AR bait 428

was used on Round Island, Mauritius. The delay in mortality was presumed to be a result of 429

some physiological difference between reptiles and bird and mammals (Merton, 1987). If 430

some reptiles are susceptible to AR poisoning but exhibit substantially delayed mortality, 431

they may be extremely effective vectors to vulnerable species in higher trophic levels if they 432

are able to ingest higher levels of rodenticide over the pre-lethal period and if mortality is 433

preceded by behaviours which increase the likelihood of predation. Laboratory toxicity tests 434

are needed across a representative suite of reptile taxa to resolve questions around the 435

dangers posed to reptiles by ARs and the capacity of reptiles to vector ARs to higher trophic 436

levels. Extensive testing of wild reptiles would be useful in assessing exposure rates and 437

ecological impacts of reptile exposure to ARs. 438

Primary consumption of ARs by reptiles through direct consumption of baits 439

intended for rodents also requires additional evaluation as a source of AR contamination in 440

terrestrial ecosystems. In captive trials, some but not all skinks (Oligosoma maccanni) 441

consumed or licked pindone bait, with increased consumption when the bait was wet 442

(Freeman et al., 1996). Direct consumption of brodifacoum baits by Shore Skinks 443

(Oligosoma smithi) in the wild has been observed in New Zealand (Wedding et al., 2010). 444

Wedding et al. (2010) cite records of five other skink species eating cereal baits, some of 445

which contained rodenticides. Bennison et al. 2016 used dye tracers to prove that the large 446

carnivorous King’s Skink (Egernia kingii) had ingested non-toxic baits laid out on islands off 447

the West Australian coast. King’s Skinks were subsequently observed consuming baits 448

containing brodifacoum during the course of a rat eradication on Penguin Island in Western 449

Australia, despite the use of specially designed bait containers intended to exclude the 450

skinks (Bettink, 2015). Others have observed bobtails (Tiliqua rugosa) – another large 451

36

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

communication). 453

These examples are cause for concern, as both bobtails and large skinks in the genus 454

Egernia have been documented as prey remains at Wedge-tailed Eagle (Aquila audax) nests 455

across a large geographic area (Brooker and Ridpath, 1980). In one instance, remains of 13 456

bobtails were found below a Wedge-tailed Eagle nest on a single visit (Simon Cherriman, 457

unpublished data). Wedge-tailed Eagles are important top carnivores in Australian food 458

webs and are highly susceptible to toxicity from the anticoagulant rodenticide pindone 459

relative to other bird species tested (Martin et al., 1994). Other carnivorous birds and 460

mammals with a higher proportion of reptiles in their diet could potentially be at greater 461

risk. 462

Reptiles could also potentially serve as an effective vector of ARs between 463

invertebrates which consume baits and more sensitive vertebrates at higher trophic levels. 464

Invertebrates have been implicated in directly vectoring rodenticides to bird species 465

including New Zealand Dotterels (Charadrius obscurus aquilonius) (Dowding et al., 2006) and 466

nestling Stewart Island robins (Petroica australis rakiura) (Masuda, Fisher, & Jamieson, 467

2014) as well as the insectivorous European hedgehog (Erinaceus europaeus) (Dowding et 468

al., 2010). If the relative tolerance of ARs demonstrated by Weir et al. (2015) is consistent 469

across numerous reptile taxa, the potential for reptiles to bioaccumulate and biomagnify 470

ARs from lower trophic levels and subsequently retain them for long periods of time makes 471

insectivorous reptiles a potentially important and widely unrecognised vector for 472

anticoagulant rodenticides to more susceptible fauna in higher trophic levels. 473

In Australia, some reptile species, particularly goannas (Varanus spp.), are a 474

culturally and economically important component of a traditional diet for some indigenous 475

peoples (Scelza et al., 2017). Liver tissue of varanids is consumed by some indigenous 476

groups (Caroline Long, Personal communication) and fatty tissues of monitor lizards are 477

eaten preferentially to other body parts (Gracey, 2000). Some rodenticides are known to 478

accumulate to high levels in fat tissue in mammals (Fisher et al., 2015) but accumulation 479

patterns in reptiles are unknown. During the course of a rodent eradication on islands in 480

Western Australia, bungaras (Varanus gouldii) were “observed eating dead and dying rats to 481

37

the extent that some droppings contained the green dye from the bait” which contained 482

brodifacoum but no mortalities were observed (Burbidge, 2004). These observations raise 483

concerns that if baiting has occurred in or near areas where traditional hunting of varanids 484

takes place, the consumption of varanid tissues likely to accumulate ARs may present a 485

previously unrecognised human health and safety risk. Consumption of feral cats by 486

indigenous people may pose another pathway for rodenticide exposure, as feral cats have 487

been killed by secondary AR poisoning during baiting events in New Zealand (Alterio, 1996). 488

Several studies have cautioned against the consumption of wild game in areas where ARs 489

have been used (Eisemann and Swift, 2006; Pitt et al., 2011), particularly SGARs (Eason et 490

al., 2001). The risks posed by consumption of varanids may be substantially greater than 491

risks associated with rabbit consumption for several reasons. Unlike rabbits which are 492

targeted in discrete baiting events with a FGAR for which there is an established withholding 493

period, varanids are not exposed in a predictable manner and may be chronically exposed to 494

stronger and more persistent SGARs with no established withholding period. The presumed 495

greater physiological tolerance of varanids to ARs and the regular consumption of varanid 496

livers as part of traditional practices considerably elevate the risks associated with varanid 497

consumption relative to rabbit consumption. Urgent investigation of potential rodenticide 498

accumulation in varanids is needed but should take into consideration the high value of this 499

taxon as a traditional food source and the cultural importance of traditional hunting 500

practices. Use of wild reptiles as a food resource is most common in tropical and 501

subtropical areas of the world (Klemens and Thorbjarnarson, 1995) where the prevalence of 502

ARs in wildlife has not been well-studied. 503

The limited literature available suggests that some reptile species are capable of 504

direct bait consumption, long AR retention time, and a capacity to tolerate and biomagnify 505

high concentrations of potent SGARs. These attributes potentially greatly increase the risk 506

of secondary and tertiary vectoring of ARs to more susceptible bird and mammal species in 507

higher trophic levels relative to other regions of the world where small mammals are 508

believed to be the primary vectors. Additional research into the prevalence of AR exposure 509

across a representative sample of reptile taxa will be critical to evaluating the threat of 510

secondary AR poisoning to wildlife in Australia and other countries with high abundance and 511

diversity of reptiles. Depending on the severity and extent of exposure detected, additional 512

38

work may be warranted to investigate the pathways driving this exposure and the role that 513

reptiles play in vectoring rodenticides to animals in higher trophic levels including humans. 514

Conclusions and Recommendations 515

Most research on exposure of non-target wildlife to ARs has been conducted in cool 516

temperate regions, particularly in North America, Europe, and New Zealand. Patterns of 517

exposure detected in these studies may differ from those in Australia and other tropical and 518

warm arid countries due to differences in the specific ARs used, regulations governing use, 519

and fundamental differences in the taxonomic composition and susceptibility of native 520

fauna. A better understanding of existing knowledge gaps will facilitate more effective and 521

scientifically-informed mitigation measures in Australia and countries with similar climates. 522

In Australia, individuals from 37 species across different feeding guilds, trophic 523

levels, and taxonomic groups have tested positive for AR exposure or are suspected to have 524

been lethally poisoned but most documentation is anecdotal or opportunistic in nature. 525

Instances of poisoning were documented across a wide range of geographic areas but 526

spatial patterns of AR exposure are poorly understood. To date, no thorough investigations 527

directly testing for AR exposure in Australian wildlife have been conducted. Island 528

eradications, feral rabbit control, agricultural application, and residential use have all been 529

implicated as sources of ARs which caused non-target wildlife mortality but the relative 530

contributions of these sources have not been quantified. 531

In aggregate, what little research exists on the interaction between reptiles and ARs, 532

suggests that at least some reptile species may be relatively resistant to the effects but likely 533

to be exposed at high levels. Physiological tolerance, coupled with long retention times 534

could make reptiles effective vectors of ARs in areas of the world where reptiles are 535

abundant. Understanding these dynamics will be critical to understanding the ecology of 536

ARs in tropical and warm arid climates where impacts on wildlife are largely unknown. 537

Effective vectoring of ARs by reptiles poses a potential unevaluated risk to human health in 538

areas where wild reptiles are harvested for human consumption. 539

At present, Australia’s regulatory framework governing the use of ARs is not 540

consistent with emerging practices in other industrialized nations. Restricting SGARs to 541

licensed users and indoor use will likely reduce the incidence and severity of non-target 542

39

poisoning and the use of lockable bait boxes could reduce risks to children and pets. 543

Coupling these proposed changes with targeted monitoring of rodenticide residues in 544

selected sentinel species will be important in evaluating the efficacy of regulatory changes 545

at reducing non-target mortality. In areas where rodents have developed resistance to 546

FGARs, use of other classes of rodenticides with lower risk of bioaccumulation (such as 547

cholecalciferol) may be a viable option for rodent control with substantially reduced risk of 548

secondary toxicity. At minimum, greater public availability of information on the types, 549

quantities, and locations of ARs sold is necessary to evaluate the risks they pose to non-550

target wildlife and humans. 551

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

Development of species-specific exposure risk models for carnivorous and 553

omnivorous fauna based on life history parameters 554

Systematic nation-wide testing of multiple taxa of carnivorous and omnivorous 555

wildlife for AR exposure, especially: 556

o species of conservation concern 557

o species consuming small mammals and carrion 558

o marsupial carnivores and scavengers 559

o reptile carnivores and scavengers 560

Systematic long-term testing of geographically widespread and common sentinel 561

species to detect temporal and spatial patterns in AR prevalence 562

Evaluation of the relative contributions of residential, commercial and agricultural 563

use of ARs to wildlife poisoning in Australia 564

o Examine incidence of non-compliance with existing legislation governing AR 565

use 566

o Collection and evaluation of data relating to AR sales and application in 567

Australia 568

Evaluation of the net impact on biodiversity of the use of pindone in and around 569

bushland areas 570

Captive testing of the sensitivity of a wider suite of wildlife species, especially 571

marsupial carnivores and reptiles to SGARs and pindone 572

40

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

nations 574

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

576

Acknowledgements 577

This project was supported financially by The Holsworth Wildlife Research Endowment via 578

The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird Research Award, 579

the Edith Cowan University School of Science Postgraduate Student Support Award, the 580

Eastern Metropolitan Regional Council’s Healthy Wildlife Healthy Lives program, the Society 581

for the Preservation of Raptors, and Sian Mawson. We thank Allan Burbidge and three 582

anonymous reviewers for improving this manuscript. Images used in the production of the 583

graphical abstract were developed by Tracey Saxby, Jane Hawkey, and Joanna Woerner of 584

the Integration and Application Network, University of Maryland Center for Environmental 585

Science (ian.umces.edu/imagelibrary/). 586

587

41

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

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

Therapeutic Goods Administration, 2017) 590

591

Schedule 4. – Prescription Only Medicine, or Prescription Animal Remedy – Substances, the 592

use or supply of which should be by or on the order of persons permitted by State or 593

Territory legislation to prescribe and should be available from a pharmacist on prescription. 594

Schedule 5. – Caution – Substances with a low potential for causing harm, the extent of 595

which can be reduced through the use of appropriate packaging with simple warnings and 596

safety directions on the label. 597

Schedule 6. – Poison – Substances with a moderate potential for causing harm, the extent of 598

which can be reduced through the use of distinctive packaging with strong warnings and 599

safety directions on the label. 600

Schedule 7. – Dangerous Poison – Substances with a high potential for causing harm at low 601

exposure and which require special precautions during manufacture, handling or use. These 602

poisons should be available only to specialised or authorised users who have the skills 603

necessary to handle them safely. Special regulations restricting their availability, 604

possession, storage or use may apply. 605

606

607

608

42

Chapter 3 Anticoagulant rodenticide exposure in an Australian 609

predatory bird increases with proximity to developed habitat 610

611

Lohr, M. T. (2018). Anticoagulant rodenticide exposure in an Australian predatory bird 612

increases with proximity to developed habitat. Science of the Total Environment. 613

643:134–144. https://doi.org/10.1016/j.scitotenv.2018.06.207 614

615

Abstract 616

Anticoagulant rodenticides (ARs) are commonly used worldwide to control 617

commensal rodents. Second generation anticoagulant rodenticides (SGARs) are highly 618

persistent and have the potential to cause secondary poisoning in wildlife. To date no 619

comprehensive assessment has been conducted on AR residues in Australian wildlife. My 620

aim was to measure AR exposure in a common widespread owl species, the Southern 621

Boobook (Ninox boobook) using boobooks found dead or moribund in order to assess the 622

spatial distribution of this potential threat. A high percentage of boobooks were exposed 623

(72.6%) and many showed potentially dangerous levels of AR residue (>0.1mg/kg) in liver 624

tissue (50.7%). Multiple rodenticides were detected in the livers of 38.4% of boobooks 625

tested. Total liver concentration of ARs correlated positively with the proportions of 626

developed areas around points where dead boobooks were recovered and negatively with 627

proportions of agricultural and native land covers. Total AR concentration in livers 628

correlated more closely with land use type at the spatial scale of a boobook’s home range 629

than at smaller or larger spatial scales. Two rodenticides not used by the public 630

(difethialone and flocoumafen) were detected in boobooks indicating that professional use 631

of ARs contributed to secondary exposure. Multiple ARs were also detected in recent 632

fledglings, indicating probable exposure prior to fledging. Taken together, these results 633

suggest that AR exposure poses a serious threat to native predators in Australia, particularly 634

in species using urban and peri-urban areas and species with large home ranges. 635

Introduction 636

Anticoagulant rodenticides (ARs) are commonly used in residential, commercial, and 637

agricultural settings for the control of rodent pests (Rattner et al., 2014b). They block the 638

43

recycling of vitamin K in the liver, which subsequently disrupts normal blood clotting in 639

vertebrates (Park et al. 1984). ARs are often divided into first generation anticoagulant 640

rodenticides (FGARs) and second generation anticoagulant rodenticides (SGARs) based on 641

their chemical structure and when they were first synthesized. Unlike FGARS, SGARs are 642

often lethal with a single feed and are substantially more persistent in liver tissue (Erickson 643

and Urban, 2004). 644

AR exposure and subsequent mortality have been detected in non-target wildlife in 645

all parts of the world where exposure has been tested (Laakso et al., 2010). Predatory bird 646

species are particularly vulnerable to AR poisoning due to a greater susceptibility to most 647

ARs than other bird species (Herring et al., 2017) and a prey base which frequently contains 648

rodents targeted by the use of ARs. In some raptor species, mortality from AR exposure 649

may have population-level impacts (Thomas et al., 2011). Unlike in Europe and North 650

America, where the non-target impacts of ARs have been extensively studied, relatively 651

little research has been conducted on AR exposure in Australian wildlife (Lohr and Davis, 652

2018; Olsen et al., 2013). This knowledge gap exists despite several lines of evidence 653

suggesting that patterns of regulation and usage in combination with differences in faunal 654

assemblages may increase the incidence and severity of non-target AR poisoning in Australia 655

relative to better-studied areas of the world (Lohr and Davis, 2018). 656

Within Australia, patterns in the spatial distribution of AR exposure have not been 657

studied in any wildlife species. A number of studies have addressed the spatial ecology of 658

anticoagulant rodenticide exposure in non-target wildlife but have been primarily limited to 659

North American mammals. Of these, some have focused on impacts within specific habitat 660

types (Cypher et al., 2014; Gabriel et al., 2012). Studies examining patterns of AR exposure 661

between urban and rural habitats have found correlations between the use of urban habitat 662

and exposure rates in San Joaquin kit foxes (Mcmillin et al., 2008) and bobcats (Riley et al., 663

2007). A model developed to predict exposure patterns in San Joaquin kit foxes found that 664

exposure was most likely in areas of low density housing on the urban/rural interface 665

(Nogeire et al., 2015). Similar dynamics have been suggested but not tested in predatory 666

bird species. Studies in North America and Europe have noted that predatory bird species 667

which use more developed habitats tend to have greater rates of AR exposure than those 668

which predominantly use more natural landscapes (Albert et al., 2010; Christensen et al., 669

2012). Additionally, a study in Spain noted a positive correlation between human 670

44

population density and AR exposure in a sample of 11 species of predatory birds and 671

mammals (López-Perea et al., 2015). The greater use of rodenticides and higher prevalence 672

of targeted commensal rodents in human-dominated landscapes relative to natural areas is 673

likely to drive these observed and suggested differences in non-target exposure. However, 674

because AR usage patterns differ between urban and agricultural environments (Lohr and 675

Davis, 2018) a need exists to evaluate the possibility of differences in non-target exposure 676

patterns between different types of anthropogenic landscapes. 677

To address this knowledge gap, I sought to compare anticoagulant rodenticide (AR) 678

exposure across intact native bushland and two different types of anthropogenic 679

landscapes. Additionally, I undertook the first large-scale targeted testing of wildlife for AR 680

exposure in the continent of Australia (Lohr and Davis, 2018). Testing was conducted on 681

Southern Boobooks (Ninox boobook), which provide an excellent model to quantify the 682

spatial distribution of threatening processes associated with fragmentation due to their 683

presence across multiple habitat types and high abundance relative to other predatory bird 684

species. To the best of my knowledge, no studies have directly addressed the relative 685

impacts of different types of human land use on AR exposure in non-target wildlife. 686

Understanding how different types of human land use impact the likelihood of AR exposure 687

in non-target wildlife will be critical in evaluating risks to wildlife on a continental scale and 688

will enable more effective targeting of measures to mitigate secondary toxicity. 689

Methods 690

Southern Boobooks are medium-sized hawk owls found across the majority of 691

mainland Australia and adjacent parts of Indonesia and New Guinea (Olsen, 2011a). They 692

are assigned a conservation status of “Least Concern” by the IUCN (“Ninox boobook,” 2018). 693

Some taxonomies consider Southern Boobooks to be synonymous with the closely-related 694

New Zealand Morepork (Ninox novaseelandiae) found in Tasmania and New Zealand but 695

recent genetic and bioacoustic evidence suggests otherwise (Gwee et al., 2017). Boobooks 696

are dietary generalists, consuming a wide variety of vertebrate and invertebrate prey 697

(Higgins, 1999; Trost et al., 2008). These dietary habits make them an ideal model species 698

for broad assessment of contamination of food webs by persistent pollutants like ARs. Their 699

presence in most habitat types across Australia, with the exception of treeless deserts 700

45

(Higgins, 1999), facilitates examination of differences in exposure across multiple habitat 701

types and allows for future replication of this study at sites across the continent. 702

Specimen Collection 703

Dead boobooks found in Western Australia were solicited from a network of 704

volunteers, wildlife care centres, and government departments and were opportunistically 705

collected when encountered. Boobooks euthanized by veterinarians and wildlife 706

rehabilitators due to severe disease or injury were included. Dates and locations where 707

each boobook was initially collected were recorded from the collector when possible. If 708

liver tissue was identifiable and had a mass >3g, it was removed and stored frozen at 20°C 709

until analysed for AR residues. A total of 73 usable boobook livers were stored for testing. 710

While an effort was made to obtain boobooks from a diversity of geographical areas and 711

habitat types throughout Western Australia, most samples originated in the more densely 712

settled urban and peri-urban areas in the south-west of Western Australia in and around the 713

city of Perth. 714

Rodenticide Analysis 715

Liver samples were analysed by the National Measurement Institute (Melbourne, 716

Australia) for residues of three FGARs (warfarin, coumatetralyl, and pindone) and five SGARs 717

(difenacoum, bromadiolone, brodifacoum, difethialone, and flocoumafen) registered for use 718

in Australia by the Australian Pesticides and Veterinary Medicines Authority. For each 719

sample, 10ml of reverse osmosis water and one gram of liver tissue were added to a 50ml 720

analytical tube and shaken for 15 minutes on a horizontal shaker. A 10ml volume of 5% 721

formic acid in acetonitrile solution was then added and the tube was shaken for an 722

additional 30 minutes. QuEChERS extraction salt was added and the tube was shaken for an 723

additional two minutes. The tube was then centrifuged for 10 minutes at 5100rpm. After 724

pipetting 3ml of the supernatant into a 15 ml analytical tube, 5ml of hexane was added and 725

the tube was shaken for two minutes then centrifuged for 10mins at 5100rpm. The hexane 726

layer was removed using a vacuum pipette and discarded. A 1ml aliquot of the supernatant 727

was transferred to a 2ml QuEChERS dispersive tube, shaken for one minute, and centrifuged 728

at 13000rpm for three minutes. The QuEChERS supernatant was then filtered using a 729

0.45μm filter. After filtration, 3μl of coumachlor was added as an internal standard to 497μl 730

of the filtered extract and vortexed prior to LC-MS/MS analysis. A Waters TQS Tandem 731

46

Quadrupole Detector Liquid Chromatograph-Mass Spectrometer (LC-MS/MS) and an 732

Acquity UPLC CSH C18 100 x 2.1mm column were used to quantify concentrations of each 733

rodenticide. Recovery rates for each AR, were calculated using chicken liver samples spiked 734

with analytical standards (Table 3.1). 735

Table 3.1 Limit of detection (LOD), limit of quantification (LOQ), average recovery, and relative standard deviation (RSD) for 736 eight ARs in a spiked chicken liver matrix. 737

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

Statistical Analysis 739

Total AR liver concentration is commonly used to compare toxicity risk when 740

individuals are exposed to multiple rodenticides (Christensen et al., 2012) due to similarities 741

in their modes of action and likely cumulative effects (Hughes et al., 2013). For this reason, 742

the sum of all liver rodenticide concentrations above the limit of detection was calculated 743

for each individual for the purposes of comparing differences in exposure by age, season, 744

and land use. In order to compare seasonal trends in total AR concentration, boobooks 745

were assigned to four groups based on their collection date: summer (December –746

February), autumn (March-May), winter (June- August), and spring (September-November). 747

All boobooks with known collection months (n=71) were included in the seasonal analysis. 748

The Kruskal-Wallis test was used to assess whether significant differences existed in liver AR 749

concentration by season. 750

Boobooks were assigned to age classes of less than one year ("hatch year") or 751

greater than one year ("after hatch year") based on the presence of juvenile down and by 752

examination of fluorescence patterns under ultraviolet light (Weidensaul et al., 2011). In 753

one instance, it was not possible to determine age class due to degradation of porphyrins 754

caused by prolonged exposure of ventral remiges to sunlight. A total of 72 boobooks of 755

determined age class were available for analysis of the relationship between age and AR 756

exposure. I used a Mann-Whitney-Wilcoxon test to determine whether total liver 757

47

concentration of ARs varied between the two age classes. Results were considered 758

significant if p<0.05. 759

Exposure Thresholds 760

The utility of rodenticide concentration in liver tissue as a means to diagnose lethal 761

exposure has been questioned (Erickson and Urban, 2004; Thomas et al., 2011) as 762

susceptibility to acute toxicity can vary among individuals and across species (Thomas et al., 763

2011). Exposure to multiple ARs adds additional complexity to the assessment of likely 764

impacts from residual liver concentrations (Murray, 2017). However, a need exists to 765

estimate likely impacts across exposed individuals and to compare the magnitude of 766

exposure to previous studies. Accordingly, I identified relevant literature which established 767

commonly used guidelines for outcomes of various exposure rates in related taxa to allow 768

estimation of likely impacts on boobooks. 769

The Rodenticide Registrants Task Force suggested that a 0.7 mg/kg liver 770

concentration of brodifacoum was likely to be toxic based largely on captive studies of Barn 771

Owls (Kaukeinen et al., 2000), however this threshold estimate may be too high, as 772

environmental conditions affecting wild birds may increase their susceptibility to ARs 773

relative to captive birds (Mendenhall and Pank, 1980). Dowding et al. (1999) estimated a 774

lethal liver concentration for brodifacoum of 0.5 mg/kg using 29 individuals from 10 species 775

of birds. Numerous studies have reported thresholds of 0.2 mg/kg (Albert et al., 2010; 776

Christensen et al., 2012; Hughes et al., 2013; Langford et al., 2013; López-Perea et al., 2015; 777

Stansley et al., 2014; Walker et al., 2008) and 0.1 mg/kg (Albert et al., 2010; Christensen et 778

al., 2012; Langford et al., 2013; Ruiz-Suárez et al., 2014; Shore et al., 2016; Stansley et al., 779

2014; Walker et al., 2011, 2008) as indices of lower limits at which lethal AR toxicity was 780

likely to occur in predatory birds. These estimates were based on two studies examining 781

wild barn owls: Newton et al. (1999) and Newton et al. (1998) respectively. I also included a 782

threshold of 0.01mg/kg as this is the lowest published record of lethal SGAR toxicity in a 783

predatory bird species (Stone et al., 1999). Boobook liver concentrations were compared 784

against these thresholds (0.7 mg/kg, 0.5mg/kg, 0.2 mg/kg, 0.1 mg/kg, and 0.01mg/kg) to 785

facilitate a comprehensive understanding of overall potential impacts of ARs across all 786

sampled individuals. 787

48

Spatial Analysis 788

Only boobooks with accurate location data were included in the spatial analysis. In 789

one instance, two road-killed boobooks were recovered at the same location. One of these 790

was randomly removed from the spatial analysis, leaving a total of 66 boobooks available 791

for analysis. Land cover for the state of WA was classified into developed, agriculture, 792

native vegetation or open water. The developed category included all areas with 793

anthropogenic impervious surfaces (roads, buildings car parks, etc.) as well as intensive land 794

uses that did not qualify as agriculture (mines, landfills, sports grounds, golf courses etc.). 795

The agriculture category included a diversity of irrigated and dryland crops, orchards, and 796

grazed areas. Intensive indoor animal agriculture was included in the developed category 797

rather than agriculture because it consisted primarily of buildings and other impervious 798

surfaces. Areas subjected to silvicultural practices were classified as part of the native 799

vegetation category due to structural similarity. Additionally like native bushland, the only 800

anticoagulant permitted for use in forestry is pindone which is used to control rabbits in 801

areas too close to human habitation to allow the safe use of 1080. Percentages of each 802

classification were calculated within circular buffer zones (areas of influence) of three 803

different sizes around each location where a boobook was found. The two smaller buffer 804

sizes were calculated to match the mean area of a boobook’s core home range (7.3 ha) and 805

total home range (145.1 ha) (Olsen et al., 2011). The largest buffer size was an arbitrarily 806

large area with a 3km radius (2827.4 ha). This larger buffer was included to account for the 807

possibility of movement of contaminated prey into boobooks’ home ranges from adjacent 808

areas influencing the probability of boobook exposure to ARs. Because open water was not 809

considered to be usable space, the percentages of the other three habitat types were 810

calculated excluding any open water within the buffers. 811

I used general linear models with a negative binomial distribution, following 812

methodology used by Christensen et al. (2012), to analyse differences in rodenticide 813

exposure by habitat composition at the three different spatial scales. The Akaike 814

Information Criterion AIC was used to rank models for habitat proportions at each spatial 815

scale. Only single variable models were considered in the ranking due to nesting and 816

correlation of habitat proportions and spatial scales. I calculated McFadden's pseudo-R2 817

values for each habitat type and spatial scale combination. Statistical analysis was 818

performed using RStudio 1.1.383 (RStudio, Inc., Boston, MA, USA). 819

49

Results 820

While I did not directly quantify physiological signs of rodenticide poisoning due to 821

most carcasses being damaged as a result of vehicle collisions, during dissection I observed 822

symptoms associated with acute lethal AR toxicity in at least nine boobooks exhibiting no 823

sign of trauma. These symptoms included excessive bleeding from minor lacerations, pale 824

or mottled livers, subdermal and muscular haemorrhage in the absence of trauma, blood in 825

the thoracic cavity, and blood around the mouth and nares. Similar symptoms have been 826

described in association with lethal AR toxicity in other raptor species (Murray, 2017). 827

50

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

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

51

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

Species Location n Individuals % Exposed % Multiple Exposure

% >0.1 mg/kg

% >0.2 mg/kg

Mean Exposure (mg/kg) (SE)

Source

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 100 94 72 16

Shore et al., 2016

Red Kite (Milvus milvus) Scotland 114 69.3 36

17.5 0.155 (0.017) Hughes et al., 2013

Buzzard (Buteo buteo) Scotland 479 44.3 14.2

2.1 0.047 (0.004) Hughes et al., 2013

Kestrel (Falco tinnunculus) Scotland 22 40.9 17.4

9.1 0.173 (0.082) Hughes et al., 2013

Barn Owl (Tyto alba) Scotland 63 34.9 17.5

17.5 0.076 (0.018) Hughes et al., 2013

Tawny Owl (Strix aluco) Scotland 34 38.2 5.9

2.9 0.047 (0.021) Hughes et al., 2013

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 58 84 52 17.2

Walker et al., 2011

Red Kite (Milvus milvus) United Kingdom 18 94 89

Walker et al., 2011

Kestrel (Falco tinnunculus) United Kingdom 20 100 95

Walker et al., 2011

Barn Owl (Tyto alba), Barred Owl (Strix varia), and Great Horned Owl (Bubo virginianus)

Canada 164 92

32

15 0.107

Albert et al., 2010

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

52

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

Barbary Falcon (Falco pelegrinoides) Spain (Canary Islands) 16 31.2 0.0915 Ruiz-Suárez et al., 2014

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

53

All Species Denmark 430

73

Christensen et al., 2012

54

ARs were detected in 72.6% of all boobook liver samples (Table 3.2) with a mean 841

summed AR exposure of 0.310 mg/kg (SE 0.069246735) (Table 3.3). Approximately 17.8% of 842

boobook livers contained greater than the suspected lethal threshold of 0.5 mg/kg total ARs 843

(Figure 3.1) with 13.7% above the more conservative limit of 0.7 mg/kg. Seven of the ten 844

boobooks with AR liver concentrations above 0.7 mg/kg appear to have died directly of AR 845

poisoning and the other three showed signs of poisoning described by Murray (2017) 846

despite other apparent proximate causes of death. More than half of the boobooks tested 847

had liver concentrations above 0.1 mg/kg (Figure 3.1) and would likely have experienced at 848

least some degree of coagulopathy (Rattner et al., 2014a). The majority of boobooks 849

(65.8%) were exposed at a level above 0.01 mg/kg – the lowest observed lethal threshold in 850

an owl (Figure 3.1). 851

852

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

The three FGARs tested – coumatetralyl, warfarin, and pindone – were infrequently 855

detected and accounted for only 0.01% of all ARs detected (Table 3.2). Coumatetralyl and 856

pindone were not detected in any of the samples and warfarin was detected in two 857

individuals at low levels (0.0024 mg/Kg and 0.0014 mg/Kg). The lower of these was below 858

the limit of quantification. Detectable exposure to SGARs was substantially higher (Table 859

3.2). Brodifacoum – the most commonly detected SGAR – was found in 72.6% of samples 860

55

and made up 84.0% of all rodenticides detected by mg/kg. It was detected in all liver 861

samples containing AR residues (Table 3.2). Difethialone and flocoumafen, which were not 862

known to be in use by the public were also detected in boobooks. Two or more ARs were 863

detected in 38.4% of boobooks tested (Figure 3.2). A maximum of five different ARs was 864

detected in two individual boobooks. 865

866

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

Mean total liver concentration of ARs was not significantly different between age 868

classes (p= 0.34). AR exposure was greatest in boobooks collected in winter and winter 869

concentrations were significantly different from summer concentrations (p=0.026) (Figure 870

3.3). The livers of two recent fledglings still under parental care contained low but 871

quantifiable amounts of brodifacoum (0.022 and 0.051 mg/kg) and difethialone (0.020 and 872

0.022 mg/kg). 873

56

874

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

Total AR exposure was positively correlated with the amount of developed area 877

within buffers at all spatial scales (Table 3.4). Proportions of agriculture and bushland 878

habitat within buffers were negatively correlated with total AR exposure at all spatial scales 879

(Table 3.4). The three AIC top-ranked models quantified habitat composition at the scale of 880

a full boobook home range and were all statistically significant (Table 3.4). The top-ranked 881

model used developed habitat at the scale of a boobook’s total home range and was highly 882

significant (p=0.00182). Correlations between the top three ranked models and total AR 883

57

concentration were not particularly strong but are stronger than would be suggested by 884

interpretation of traditional R2 indices, as McFadden's pseudo-R2 values falling in the range 885

of 0.2 to 0.4 “represent an excellent fit” (McFadden, 1978). 886

Table 3.4 Akaike information criterion (AIC) ranking of models of the association between percentage of single land use 887 types within buffers around collection points and total anticoagulant rodenticide liver concentration in Southern Boobooks 888 (n= 66) in Western Australia at three different spatial scales (Big=2827.4 ha buffer, Mid=145.1 ha buffer, Small=7.3 ha 889 buffer. 890

Model Estimate Std. 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 Vegetation -2.5139 0.9584 -2.623 0.00871 754.35

0.05081192

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 Vegetation -1.9017 1.066 -1.784 0.0744 756.8

0.01968855

891

Discussion 892

The overall proportion of boobooks with detectable AR exposure (72.6 %) and the 893

proportion of boobooks exposed to two or more rodenticides (38.4%) was high but within 894

the range of estimates generated by studies in Europe and North America (Table 3.3). Mean 895

total AR concentration in boobooks (0.310 mg/kg) was substantially higher than any other 896

available published estimate with the exception of Red Kites (Milvus milvus) (0.413 mg/kg) 897

in Denmark (Christensen et al., 2012). The extremely high mean exposure in boobooks may 898

result from multiple causes. A large proportion of samples were obtained from urban and 899

peri-urban areas where exposure is likely to be more prevalent. This was also the case in 900

several other studies documenting high exposure rates and liver concentrations (López-901

Perea et al., 2015; Murray, 2017; Stansley et al., 2014). As a consequence, the sample of 902

boobooks used in this study is probably not representative of Australia as a whole but may 903

provide a useful estimate for other large human population centres elsewhere. Circadian 904

activity patterns may also increase boobooks’ risk of AR exposure relative to some other 905

raptor species. Nocturnal species have been noted to have higher liver AR concentrations 906

58

than diurnal species (Ruiz-Suárez et al., 2014; Sánchez-Barbudo et al., 2012). If owls using 907

highly populated landscapes are at greater risk than other bird species, future evaluation of 908

Powerful Owls which use urban and peri-urban areas and are listed as vulnerable in Victoria 909

may be warranted. Southwest populations of Masked Owls (Tyto novaehollandiae) and 910

Barking Owls (Ninox connivens), both of which are listed as P3 priority fauna (poorly known 911

but thought to be possibly threatened) in Western Australia, may also be susceptible to AR 912

poisoning in areas where developed habitats are encroaching on their remaining ranges. 913

As a consequence of the methodology used in sample collection, this study probably 914

underestimates the proportion of lethal poisonings which actually occur. Anticoagulant 915

rodenticides induce lethargy prior to mortality and lethally poisoned owls are more likely to 916

die in nest hollows or roost sites in dense vegetation where their likelihood of detection by 917

humans would be low (Newton et al., 1990). Similar underestimation of lethal toxicity has 918

been suggested in studies of mammals exposed to ARs, as well (Mcdonald et al., 1998). 919

Conversely, if haemorrhaging induced by sub-lethal exposure reduced a boobook’s reaction 920

time or ability to fly, it could increase the risk of other proximate sources of mortality 921

(Newton et al., 1990) such as collisions with vehicles or windows. This could potentially 922

increase its likelihood of being killed in a conspicuous location and subsequently collected 923

for this study with the end result of inflating the number of sub-lethally exposed birds 924

entering this study. 925

Individual Rodenticides 926

A lack of detectable pindone residues in the livers of the boobooks sampled was 927

unexpected because pindone is used within the Perth metropolitan area to control rabbits 928

in urban bushlands and previous literature implicates similar control programs elsewhere in 929

Australia in secondary poisonings of native raptors (Olsen et al., 2013) though this has 930

recently been disputed (Olsen and Rae, 2017). Failure to detect pindone could be the result 931

of a short retention time relative to more persistent SGARs (Fisher et al., 2003), its use in 932

targeted and short-term control efforts, low overall usage relative to commercial and 933

residential use of other anticoagulant rodenticides, or dietary patterns of boobooks 934

precluding consumption of European rabbits (Oryctolagus cuniculus) – the species targeted 935

by pindone applications. While it is possible that occasional localised exposure may occur, it 936

appears that pindone, as currently applied in urban and peri-urban areas does not 937

constitute a substantial threat to boobook populations relative to other rodenticides 938

59

originating from commercial and residential sources. Future studies on impacts of pindone 939

on native raptors should consider testing species which are more likely to prey on rabbits 940

(Wedge-tailed Eagles (Aquila audax) and Little Eagles (Hieraaetus morphnoides)) (Olsen et 941

al., 2006) or scavenge rabbit carcasses (Whistling Kites (Haliastur sphenurus)) (Fuentes et al., 942

2005) and are at greater risk of secondary exposure. 943

Failure to detect coumatetralyl in any samples and the detection of warfarin at 944

extremely low concentration in only two samples despite commercial availability to the 945

public suggests that their relatively short half-life in liver tissue (Fisher et al., 2003) probably 946

reduces the incidence and severity of secondary exposure and precludes bioaccumulation 947

and biomagnification. This result is consistent with absence or low concentration and 948

prevalence of FGARs relative to SGARs in other wildlife species since SGARs came into 949

widespread use (Albert et al., 2010; Fourel et al., 2018; Murray, 2017; Ruiz-Suárez et al., 950

2014). 951

The detection of brodifacoum at rates an order of magnitude higher than all other 952

ARs combined is probably attributable to a combination of its greater duration of 953

persistence in liver tissue (Horak et al., 2018), more prevalent use, and incorporation into a 954

greater number of commercially available rodenticide bait products. This is particularly 955

concerning because captive studies suggest that brodifacoum is more likely to cause 956

secondary toxicity in birds than any other tested ARs due to its high toxicity and long liver 957

retention time (Erickson and Urban, 2004). Bromadiolone and difenacoum respectively, 958

were the next most commonly detected in samples (Table 3.2). This is probably because, 959

together with brodifacoum, they comprise the three SGARs commonly available in WA at 960

retail stores. At present, brodifacoum, bromadiolone, and difenacoum probably pose the 961

greatest threat of secondary poisoning to non-target wildlife of all ARs in use. 962

The detection of flocoumafen and difethialone – which are not readily available to 963

the public due to sale in bulk quantities but are used by pest control professionals – 964

indicates that at least some proportion of wildlife exposure is directly related to commercial 965

pest control activities. Flocoumafen was the most prevalent rodenticide detected in liver 966

tissue of one boobook, which died shortly after admission to a wildlife care centre and 967

showed physiological signs of AR poisoning (pale mottled liver, subcutaneous haemorrhage, 968

and large quantities of blood in the abdominal cavity). These findings have potentially 969

serious implications for legislation attempting to curtail non-target exposure by limiting 970

60

public access to SGARs. In the United States, legislation restricting the use of SGARs to 971

licensed professionals went into effect in 2011 (Bradbury, 2008). However, a subsequent 972

study found an increase in AR exposure in four predatory bird species in Massachusetts, 973

USA following the ban (86% of 161 birds from 2006 - 2010 compared to 96% of 94 birds 974

exposed from 2012 - 2106) perhaps due to an increased use of professional rodent control 975

services (Murray, 2017). My findings provide additional evidence that use of ARs by 976

professional pesticide applicators does contribute, at least to some degree, to poisoning of 977

non-target raptors. However, the impacts of this source relative to private use are difficult 978

to assess because other SGARs which are available to the public – particularly brodifacoum 979

and bromadiolone – are in common use by professional pesticide applicators in WA. Taken 980

together, these results cast doubt on whether regulations restricting sale of SGARs from 981

private use will be sufficient to reduce widespread exposure and toxicity in predatory birds. 982

After the completion of this study, it was brought to my attention that diphacinone 983

was also being used in Western Australia by commercial pesticide applicators. This FGAR 984

has a relatively short half-life of three days in rat liver tissue and as a consequence is 985

unlikely to bioaccumulate and cause secondary poisoning in predatory non-target wildlife 986

(Fisher et al., 2003). The registration of diphacinone in Australia has expired. However, if 987

diphacinone is re-registered, future monitoring projects should include diphacinone testing 988

as it could potentially contribute to overall rodenticide exposure. 989

Exposure to multiple rodenticides (38.4%) was relatively common in sampled 990

boobooks but not as frequent as in some other predatory bird species (Christensen et al., 991

2012; Murray, 2017; Walker et al., 2011). The relatively high rate of multiple exposure and 992

the presence of detectable levels of up to five different ARs in liver tissue suggests 993

cumulative exposure from multiple prey items over an extended period of time. This 994

hypothesis is supported by the finding that livers of adult raptors in Denmark contained 995

multiple rodenticides more frequently than those of juveniles (Christensen et al., 2012). The 996

prevalence of multiple exposures in boobooks is particularly concerning because laboratory 997

studies on rats determined that warfarin sensitivity is increased after sub-lethal exposure to 998

brodifacoum (Mosterd and Thijssen, 1991). If ARs have a synergistic effect rather than a 999

purely additive effect, raptors may be negatively impacted at a lower threshold when 1000

exposed to more than one AR, leading to underestimates of negative impacts on non-target 1001

wildlife. 1002

61

Rodenticide Thresholds 1003

The utility of detectable rodenticide concentration in liver tissue as a means to 1004

diagnose lethal exposure has been questioned (Erickson and Urban, 2004; Thomas et al., 1005

2011) as susceptibility to acute toxicity can vary among individuals and across species 1006

(Erickson and Urban, 2004). However, it can be informative in comparing environmental 1007

exposure and as an index for potential impacts at the population level. Depending on the 1008

threshold used (0.7 mg/kg or 0.5 mg/kg), either 13.7% or 17.8% of boobooks tested had 1009

rates of exposure consistent with likely lethal outcomes. Confirmation of physical signs of 1010

rodenticide poisoning in all boobooks with AR liver concentrations above 0.7 mg/kg and the 1011

absence of other obvious causes of death in 70% of these individuals indicates that this 1012

threshold is a reasonable guideline for estimating likely lethal toxicity in boobooks. 1013

Regardless of the threshold used, the relatively high frequency of exposure at levels likely to 1014

be directly lethal is cause for concern. In combination with visible signs of AR poisoning, it 1015

indicates that exposure to ARs contributed substantially to mortality in bobooks found dead 1016

or brought to wildlife carers in the urban and peri-urban areas where most samples were 1017

collected. 1018

Exposure at potentially dangerous but not necessarily lethal levels was also high 1019

relative to most published studies examining rodenticide exposure in wild raptors found 1020

dead or moribund. The proportion of boobooks exposed at levels above 0.2 mg/kg (35.6%) 1021

was higher than all other reported estimates except for in Barn Owls (Tyto alba) (57.9%) and 1022

Eagle Owls (Bubo bubo) (64.3%) in Spain (López-Perea et al., 2015) and Red Kites (Milvus 1023

milvus) (66.7%) in Denmark (Christensen et al., 2012). In all three species, the sample size 1024

was small (n<20). The percentage of boobooks with total AR liver concentrations above 1025

0.1mg/kg (50.7%) was substantially greater than all previously reported species except for 1026

Red-tailed Hawks in New Jersey, USA (47%) (Stansley et al., 2014). At minimum, a 1027

threshold of 0.1 mg/kg should be considered potentially dangerous. In a laboratory study 1028

using Eastern Screech Owls (Megascops asio), diphacinone concentrations of ≥0.1 mg/kg in 1029

liver tissue were associated with coagulopathy (Rattner et al., 2014a). Coagulopathy is likely 1030

more dangerous to wild birds due to greater amounts of movement and injuries associated 1031

with capturing prey and may have synergistic interactions with environmental stressors 1032

which increase the chance of mortality (Erickson and Urban, 2004). SGARs are also more 1033

62

toxic than diphacinone and can logically be expected to have at least as great of an impact 1034

at the same threshold. 1035

Sub-lethal exposure was common in boobooks regardless of the chosen threshold. 1036

The sub-lethal impacts of chronic AR exposure are poorly studied in wildlife. A number of 1037

lines of evidence suggest that even exposure below the threshold needed to cause lethal 1038

haemorrhage is not benign. While Thomas et al. (2011) take issue with the uncritical use of 1039

liver concentrations to assess likely toxicity, their probabilistic methodology examining AR 1040

toxicity in four raptor species predicted that 20% of individuals would experience 1041

quantifiable toxicity at levels as low as 0.08 mg/kg. Increased rates of parasitism and 1042

infectious disease have also been documented in association with AR exposure in bobcats 1043

(Lynx rufus) (Riley et al., 2007), Great Bustards (Otis tarda) (Lemus et al., 2011), and 1044

common voles (Microtus arvalis) (Vidal et al., 2009). In bobcats, immunosuppression and 1045

inflammatory response associated with chronic sub-lethal AR exposure and use of urban 1046

habitats may have led to an outbreak of notoedric mange (Serieys et al., 2018). Similar 1047

disruption of immune system function may occur in other chronically-exposed wildlife 1048

(Serieys et al., 2018). Several studies have also suggested the possibility of increased 1049

mortality rates via accidents, predation, vehicle collisions, nutritional stress, and blood loss 1050

following minor injury in wildlife exposed to sub-lethal doses of anticoagulant rodenticides 1051

(Albert et al., 2010; Mendenhall and Pank, 1980; Newton et al., 1990; Stone et al., 2003, 1052

1999). If this dynamic is indeed consistent across wildlife species, the high rates of 1053

presumably sub-lethal exposure detected in boobooks are cause for concern. If sub-lethal 1054

exposure to ARs substantially increases the risk of parasitism and other sources of mortality, 1055

it is not appropriate to assess the overall impacts of anticoagulants on predatory bird 1056

populations based solely on documentation of direct lethal toxicity. 1057

Spatial Correlations 1058

We observed weak but statistically significant correlations between AR exposure and 1059

habitat proportions in proximity to recovered boobook carcasses. The difference in the 1060

direction of correlations between AR exposure and proportions of agricultural and 1061

developed habitats, the consistency of the trends at different spatial scales, and the 1062

increasing strength of the trends at the most biologically meaningful spatial scale all suggest 1063

an actual difference in exposure risk between the two anthropogenic landscapes. Future 1064

studies on this topic should attempt to improve sample collection across different types of 1065

63

anthropogenic landscapes or focus on species for which samples are more readily available 1066

across study areas. A low sample size of boobook carcasses from landscapes predominantly 1067

comprised of native bushland or agriculture likely contributed to the low predictive value of 1068

top models. 1069

The three top-ranked models for boobook AR exposure used habitat data at the 1070

scale of an average home range. Foraging behaviour likely explains the closer correlation of 1071

AR exposure and habitat type at the spatial scale of an average boobook home range 1072

relative to other spatial scales. The vast majority of foraging occurs within an animal’s home 1073

range and its exposure to ARs can be expected to relate most closely to the proportions of 1074

habitat types likely to be sources of contamination of its prey base at this spatial scale. 1075

Boobooks have relatively small home ranges in comparison to other Australian owl species 1076

(Kavanagh and Murray, 1996; Soderquist and Gibbons, 2007). If risk of rodenticide exposure 1077

is related to developed area at the scale of an animal’s home range, species with larger 1078

home ranges may be exposed over a broader portion of the landscape. This hypothesis is 1079

supported by the finding that in bobcats – a species with a much larger home range than 1080

boobooks– the concentration but not the presence of ARs in liver tissue correlated with the 1081

proportion of developed habitat within their home range (Riley et al., 2007). Taken in 1082

combination, these results suggest that species with large home ranges are likely to be at 1083

risk of some degree of AR exposure if their home range encompasses even small areas of 1084

developed habitat. As a consequence, encroachment of human structures into large areas 1085

of natural habitat may have an impact on predatory species with large home ranges that is 1086

disproportionate to the area of habitat lost through development. 1087

The positive correlation between total AR exposure and the proportion of developed 1088

area within buffers was expected due to the widespread use of rodenticides in commercial 1089

and residential settings. This pattern of exposure has been suggested following detection of 1090

high exposure rates in densely populated areas (López-Perea et al., 2015; Stansley et al., 1091

2014) but, this appears to be the first instance where differences in exposure across habitat 1092

types has been directly quantified in a bird species. A number of other studies have 1093

examined the spatial patterns of AR exposure in wildlife. The trend in boobooks was similar 1094

to the correlation between developed areas and total AR exposure observed in a study of 1095

bobcats and mountain lions in California (Riley et al., 2007). Similarly, AR exposure was 1096

common (87%) in an urban population of San Joaquin kit foxes but no rodenticides were 1097

64

detected in individuals from a non-urban population (Mcmillin et al., 2008). Frequent AR 1098

exposure in wildlife inhabiting developed habitats is typically attributed to the “prevalent 1099

and wide-spread” use of ARs in urban areas (Cypher et al., 2014). Higher prevalence of 1100

commensal rodents which serve as vectors of ARs in urban areas may exacerbate this 1101

problem. A study in Canada demonstrated a higher proportion of rats in the diet of Barn 1102

Owls with territories containing more urban land use (Hindmarch and Elliott, 2014). 1103

Assuming that commensal rodents are an important vector of ARs, their higher relative 1104

proportion in the diets of urban owls may increase the incidence and severity of AR 1105

exposure. Boobooks are likely to be affected by this dynamic. In Canberra, Australia, 1106

boobook diets contained a higher percentage of mammal biomass in suburban areas 1107

(65.8%) than in woodland areas (26.0%) (Trost et al., 2008). Both the high prevalence of 1108

rodenticide use and the greater availability of potentially exposed commensal rodents likely 1109

contribute to the positive correlation between rodenticide exposure and developed habitat 1110

observed in boobooks. 1111

A negative correlation between AR exposure and the proportion of bushland area 1112

within simulated home ranges was expected because rodenticides are seldom used in native 1113

habitats, aside from the use of pindone to control rabbits. Only one other study has tested 1114

spatial patterns of AR exposure in wildlife primarily using bushland habitats. Unlike patterns 1115

observed in boobooks, high exposure rates were unexpectedly detected in fishers (Martes 1116

pennanti) throughout areas of forested habitat, probably as a result of rodenticide use 1117

associated with illegal marijuana production (Gabriel et al., 2012). Similarly a threatened 1118

Spotted Owl (Strix occidentalis) with illegal marijuana cultivation within its home range was 1119

documented to have been exposed to brodifacoum despite being in a remote natural area 1120

(Franklin et al., 2018). Conservation and law enforcement professionals should be aware of 1121

this potential source of environmental contamination when attempting to mitigate damage 1122

caused by illegal marijuana cultivation in remote areas in Australia. Future work examining 1123

the distance rodenticides travel into bushland ecosystems from adjacent sources will be 1124

useful in gaining a better understanding of the relationship between fragmentation and 1125

rodenticide use. This could potentially lead to establishing appropriate sizes for SGAR 1126

exclusion zones around bushland areas containing sensitive fauna and reduce edge effects 1127

relating to SGARs. 1128

65

The negative correlation between total AR exposure and the proportion of 1129

agricultural area within simulated home ranges was somewhat surprising, as rodenticides 1130

are known to be used in agricultural settings. AR exposure in wildlife has been attributed to 1131

agricultural application of ARs in the UK (Birks 1998; Hughes et al. 2013), Spain (Lemus et al., 1132

2011), France (Fourel et al., 2018), and Australia (Young and De Lai, 1997). Anecdotal 1133

accounts from farmers indicate that a variety of first and second generation products are 1134

used for asset protection around buildings and in grain storage areas in Wwestern Australia 1135

(Don Thompson personal communication). However, they are not licensed for use directly 1136

in crops or along crop perimeters. As a consequence, the total amount of bait deployed per 1137

unit area is likely to be substantially lower than in developed areas. However, in agricultural 1138

systems, total compliance with best practice application methods for SGARs may be rare 1139

and lack of compliance probably facilitates greater risk of secondary toxicity to native 1140

wildlife (Tosh et al., 2011). An anecdotal report of farmers in Western Australia requesting 1141

the FGAR pindone to control kangaroos (Twigg et al., 1999) – a use not allowed by the 1142

labelling – suggests that illegal use of ARs in agricultural contexts may be an issue in some 1143

areas. The widespread availability of SGARs to the public in Australia increases the risk that 1144

misuse could lead to localised impacts on non-target wildlife. 1145

The negative correlation between proximity to agricultural land and AR exposure 1146

may not be consistent throughout all Australian agricultural systems. In Queensland, 1147

declines in breeding owl abundance were attributed to broad-scale application of a 1148

brodifacoum-based rodenticide in canefields (Young & De Lai 1997) but this product was 1149

subsequently removed from the market (Twigg et al., 1999). At present, brodifacoum is 1150

only registered for use in and around buildings in Australia (McLeod & Saunders 2013) but 1151

can be freely purchased and applied without a license. While less toxic and persistent than 1152

brodifacoum, a coumatetralyl-based product is currently licensed for use in sugar cane, 1153

pineapple, and macadamia crops across Australia (Australian Pesticides and Veterinary 1154

Medicines Authority, 2017b). More concerningly, during rodent plagues the SGAR 1155

bromadiolone has been used to bait field perimeters in New South Wales (New South Wales 1156

Department of Primary Industries, 2011; New South Wales Government: Department of 1157

Primary Industries, 2017). 1158

66

Seasonal Differences 1159

The difference in AR exposure observed between boobook carcasses recovered in 1160

winter and those recovered in summer potentially reflects increased risk of exposure during 1161

winter when rodents make up a larger proportion of the diet. Boobooks are dietary 1162

generalists and one study indicates that boobook diet varies seasonally and includes higher 1163

proportions of vertebrates in winter than in autumn (Trost et al., 2008). This seasonal 1164

variation in diet may reduce the risk of accumulating lethal levels of ARs in boobooks 1165

relative to some other raptor species. Species preying predominantly on small mammals 1166

are likely to be at greater risk of exposure than species that prey predominantly on birds 1167

(Ruiz-Suárez et al., 2014). This hypothesis is supported by a lack of seasonal variation in AR 1168

exposure in Tawny Owls (Strix aluco) which feed consistently on bank voles (Myodes 1169

glareolus) and field mice (Apodemus spp.) (Walker et al., 2008). Similarly, in the United 1170

States, rodenticide exposure rates and concentrations did not vary significantly by season in 1171

Red-tailed Hawks (Buteo jamaicensis) (Stansley et al., 2014) which feed predominantly on 1172

mammals year-round. The only other study detecting seasonal variation in liver AR 1173

concentration found a significant difference in only one of five ARs tested (Christensen et 1174

al., 2012). This difference was attributed to an influx in autumn of migratory raptors from 1175

more sparsely populated regions with presumably less AR exposure risk (Christensen et al., 1176

2012). 1177

It is possible that consuming few rodents during a portion of the year allows 1178

boobooks to excrete sufficient levels of highly-persistent SGARs that total liver 1179

concentrations are less likely to accumulate to a lethal level. In this scenario, other raptor 1180

species which consistently consume rodents throughout the year – such as Masked Owls 1181

and Barking Owls – may be at elevated risk of lethal poisoning relative to boobooks. 1182

Alternately, seasonal variation in rodenticide exposure in boobooks could be correlated with 1183

seasonal differences in rodenticide use patterns. Information on rodenticide sales is not 1184

publicly available, but anecdotal accounts from some Perth residents indicate greater use of 1185

rodenticides in winter in response to greater perceived abundance of commensal rodents. 1186

Improved knowledge of rodenticide application patterns and seasonal patterns of 1187

rodenticide exposure in species with a more consistent mammal-based diet would be useful 1188

in addressing these questions. 1189

67

The high AR exposure rates observed in boobooks despite seasonal variation in the 1190

proportion of rodents in their diet highlights the need for additional study of exposure rates 1191

of other taxa which may potentially vector rodenticides. Documented exposure in raptors 1192

which prey primarily on birds indicates that non-rodent vectors may substantially contribute 1193

to AR exposure at higher trophic levels (Thomas et al., 2011). Invertebrates have been 1194

implicated in vectoring lethal levels of rodenticides to bird species including New Zealand 1195

Dotterels (Charadrius obscurus aquilonius) (Dowding et al., 2006) and nestling Stewart 1196

Island robins (Petroica australis rakiura) (Masuda et al., 2014) as well as an insectivorous 1197

mammal, the European hedgehog (Erinaceus europaeus) (Dowding et al., 2010). Reptiles 1198

could potentially also be effective vectors to higher trophic levels (Lohr and Davis, 2018). 1199

Further investigation of AR residues across more taxa is necessary to fully understand 1200

ecosystem-wide AR contamination and the vectors by which carnivorous species are 1201

exposed. 1202

Rodenticide in fledglings 1203

The detection of SGAR exposure in recent fledglings provides a possible indication as 1204

to why there was no significant difference in total AR exposure between hatch year 1205

boobooks and older adults. AR exposure prior to leaving the nest is particularly concerning 1206

from a conservation perspective. Suspected brodifacoum poisoning was previously 1207

reported as the likely cause of death of Norfolk Island Boobook chicks which were still in the 1208

nest (Debus, 2012) but there was no indication of physical examination or direct testing for 1209

AR exposure. Birds with growing feathers may be at additional risk of exsanguination 1210

(Newton et al., 1990). This may put chicks and recent fledglings at greater risk than adult 1211

birds which do not typically moult large proportions of their feathers simultaneously. 1212

Additional sub-lethal threats to chicks have also been reported. Stunted growth across 1213

several biometric measurements of nestling Barn Owls was observed in plots treated with 1214

anticoagulant rodenticides relative to control plots in Indonesia (Naim et al., 2010). While 1215

reduced prey availability due to rodent control likely had a negative influence on growth 1216

rates, nestlings in areas treated with the SGAR brodifacoum showed reduced growth when 1217

compared to areas where rodents were controlled with the FGAR warfarin or a biological 1218

rodent control agent (Naim et al., 2010), suggesting that AR exposure contributed to 1219

reduced nestling growth. Similarly, a dramatic reduction in breeding success occurred in a 1220

population of closely-related moreporks on Mokoia Island in New Zealand in the breeding 1221

68

season immediately following a broad-scale distribution of brodifacoum as part of an 1222

attempted mouse eradication (Stephenson et al. 1999). While Stephenson et al. (1999) 1223

concede that the reduction in breeding success may have been related to a drop in prey 1224

availability rather than a direct effect of rodenticide toxicity, depression of breeding success 1225

by anticoagulant rodenticides is plausible. Laboratory testing also detected modest 1226

reductions in weight gain and wing growth in juvenile Japanese Quail (Coturnix coturnix 1227

japonica) exposed to sub-lethal doses of brodifacoum or difenacoum (Butler, 2010). 1228

Perhaps the most conclusive evidence of negative impacts of sub-lethal AR exposure on 1229

growing birds is the correlation observed between concentrations of bromadiolone in blood 1230

and reduced body condition observed in nestling Common Kestrels (Falco tinnunculus) 1231

(Martínez-Padilla et al., 2016). 1232

Nest success may also be impacted in the early stages of nesting. Embryo toxicity 1233

has been observed in domestic chicken eggs injected with the anticoagulant rodenticide 1234

flocoumafen (Khalifa et al., 1992). It is also possible that exposure to anticoagulant 1235

rodenticides could impact egg viability via reductions in the integrity of eggshells. Exposure 1236

to therapeutic anticoagulants has resulted in bone density loss in humans by disruption of 1237

the vitamin K cycle and resultant suppression of calcification (Fiore et al. 1990; Resch et al. 1238

1991; Monreal et al. 1991) though similar effects on bone density have not been observed 1239

in birds (Knopper et al., 2007). Residues of bromadiolone and chlorophacinone were 1240

detected in yolk and albumin of addled Barn Owl eggs in areas of palm plantations treated 1241

with rodenticides but no changes to eggshell thickness or morphology were detected (Salim 1242

et al., 2015). However, changes to barn owl egg morphology, reduced eggshell mass and 1243

decreased eggshell thickness have been observed when eggs contained higher 1244

concentrations of brodifacoum (Naim et al., 2012). While teratogenic effects of 1245

anticoagulant rodenticides are not widely reported in birds, one study suggested this 1246

possibility when the authors detected a single barn owl nestling in a plot treated with 1247

brodifacoum which failed to grow primary feathers and would have been unable to fly 1248

(Naim et al., 2010). Haemorrhage of oviducts in association with rodenticide poisoning has 1249

been observed in female raptors carrying eggs (Murray, 2017), suggesting that ARs may 1250

pose a particular risk to nesting females. Future assessments of population-level impacts of 1251

anticoagulant rodenticide exposure need to consider not only adult mortality, but also 1252

impacts on fecundity and recruitment. 1253

69

1254

Conclusion 1255

My hypothesis that total AR exposure would vary between areas predominated by 1256

different types of anthropogenic landscape is to some degree supported by the finding of 1257

significant, though weak, relationships trending in opposite directions between total liver AR 1258

concentration and proportions of agriculture and developed land at the spatial scale of a 1259

boobook’s home range. Understanding this dynamic is key to assessing landscape-level risk 1260

of AR poisoning across carnivores and scavengers in Australia. It will also facilitate future 1261

attempts to model exposure risk in endangered and priority taxa which may be susceptible 1262

and will enable more specific risk assessment prior to proposed future developments. The 1263

high rates and magnitude of AR exposure raise serious concerns about AR exposure in other 1264

Australian species. Future work should evaluate the impact of ARs on other Australian 1265

wildlife, particularly species utilizing urban and peri-urban areas, species with large home 1266

ranges, and species regularly consuming commensal rodents. The detection in boobooks of 1267

ARs presumed to be used only by professionals is concerning. Ongoing review of the 1268

registration of SGARs by the APVMA should take this into consideration when evaluating the 1269

efficacy of restricting SGARs to licensed pesticide applicators in reducing poisoning in non-1270

target wildlife. 1271

1272

Acknowledgements 1273

This project was supported financially by The Holsworth Wildlife Research 1274

Endowment via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird 1275

Research Award, the Edith Cowan University School of Science Postgraduate Student 1276

Support Award, the Eastern Metropolitan Regional Council’s Healthy Wildlife Healthy Lives 1277

program, the Society for the Preservation of Raptors, and Sian Mawson. 1278

I would like to extend my appreciation to the staff of National Measurement 1279

Institute’s (NMI), Analytical Services Port Melbourne branch, in particular Hao Nguyen and 1280

her veterinary drug residue measurement team for their support in providing high quality 1281

measurements using NATA accredited LC-MSMS measurement techniques. 1282

Cheryl Lohr provided valuable assistance in statistical analysis and Shaun Molloy 1283

graciously volunteered time to help develop necessary GIS layers. I would especially like to 1284

70

thank Kanyana Wildlife Rehabilitation, Native Animal Rescue, Native ARC, and Nature 1285

Conservation Margaret River Region and the many other individuals especially Simon 1286

Cherriman, Angela Febey, Amanda Payne, and Stuart Payne for contributing samples. This 1287

manuscript was improved by comments from Dr. Robert Davis, Dr. Allan Burbidge and two 1288

anonymous reviewers. The photograph in the graphical abstract was provided by Simon 1289

Cherriman. 1290

1291

71

Chapter 4 Widespread genetic connectivity in Australia’s most 1292

common owl, despite extensive habitat fragmentation 1293

Abstract 1294

1295

Lohr, M. T., P. B. S. Spencer, S. Krauss, J. Anthony, A. H. Burbidge, and R. A. Davis. 1296

Widespread genetic connectivity in Australia’s most common owl, despite extensive 1297

habitat fragmentation. The Condor: Ornithological Applications. (In Preparation). 1298

1299

Reductions in genetic diversity and genetic connectivity have been documented in 1300

some predatory bird species in response to anthropogenic habitat fragmentation. The 1301

Australian Boobook (Ninox boobook) is the most common and widely-distributed owl in 1302

Australia but declines in abundance have been observed across its range. To investigate 1303

whether genetic factors associated with habitat fragmentation have been associated with 1304

this reduction in abundance, we used polymorphic microsatellite loci to investigate patterns 1305

of genetic variation and its spatial structure in boobooks from a variety of fragmented and 1306

relatively undisturbed landscape types across Western Australia. The maximum distance 1307

between samples was 1391 km. Genetic analysis was informed by data on post-breeding 1308

dispersal of juvenile boobooks gathered from banding data resulting from this and other 1309

studies across Australia. We found weak spatial genetic structuring and no evidence of 1310

genetic erosion associated with inbreeding in heavily fragmented landscapes. Bayesian 1311

modelling and principal coordinates analysis suggested a single large panmictic population 1312

across all areas sampled. Within the heavily fragmented landscape of an extensive urban 1313

area, band re-sightings and recoveries substantiate the considerable capacity of juvenile 1314

boobooks to disperse across areas far greater than the distance between patches. We 1315

hypothesise that the genetic homogeneity observed is a consequence of long distance 1316

dispersal capacity of boobook offspring and their ability as habitat and dietary generalists to 1317

make use of highly altered habitats. 1318

72

Introduction 1319

Habitat Fragmentation, Connectivity, and Genetic Structure 1320

Habitat fragmentation can directly negatively impact the genetic diversity of a 1321

population of organisms by restricting gene flow between habitat patches and reducing 1322

effective population size (Aguilar et al., 2008). Reduction of effective population size can 1323

increase additional risks to small populations posed by demographic stochasticity, genetic 1324

drift, and inbreeding depression (Soulé and Simberloff, 1986). In fragmented environments, 1325

highly mobile species are less likely to experience these phenomena than species with 1326

limited dispersal capacity because of greater gene flow between fragments (Bohonak, 1327

1999). However, differences between the types and severity of threats found within distinct 1328

types of habitat matrix may drive differences in matrix permeability, irrespective of the 1329

biogeographic variables and mobility of the species concerned (Collinge, 1996). 1330

Simultaneous investigation of genetic connectivity in a single species across multiple matrix 1331

types has the potential to inform our understanding of the impacts of different matrices on 1332

permeability and metapopulation dynamics. 1333

1334

Genetic Responses of Predatory Birds to Fragmentation 1335

Predators are more frequently extirpated as a result of fragmentation than animals 1336

at lower trophic levels, as a result of their larger home range requirements and smaller 1337

population sizes (Didham et al. 1998; Gilbert et al. 1998; Duffy 2003). Predatory birds 1338

specifically have been observed to be at greater risk of extinction as a result of habitat 1339

fragmentation than other bird species (Leck 1979; Brash 1987; Carrete et al. 2009). This 1340

relative sensitivity to fragmentation makes predatory birds useful bio-indicators of 1341

ecosystem health in fragmented landscapes (Rodríguez-Estrella et al., 1998). A variety of 1342

negative impacts on predatory bird populations have been documented in association with 1343

use of highly fragmented landscapes, with some differences noted between urban and 1344

agricultural landscapes. In urban landscapes, documented negative impacts include 1345

increased mortality associated with electrocutions and collisions with vehicles and 1346

anthropogenic structures (Hager, 2009), reduced nest success due to higher rates of 1347

parasitic infection (Boal and Mannan, 1999), and higher rates of exposure to anticoagulant 1348

rodenticides (Lohr, 2018). Likewise, agricultural intensification has led to declines in 1349

73

carnivorous bird abundance as a consequence of loss of nesting sites, pesticide poisoning, 1350

and overgrazing of prey species habitat (Newton, 2004) as well as continental-scale decline 1351

across farmland bird species generally (Donald et al. 2001). While differences in landscape 1352

structure and the threatening processes associated with the urban and agricultural matrix 1353

clearly differ, it is unclear how or whether the different pressures exerted on predatory 1354

birds existing in these landscape types drive differences in matrix permeability and genetic 1355

connectivity. 1356

Relatively few studies have been conducted on the genetic impacts of habitat 1357

fragmentation on predatory birds, particularly across different types of anthropogenic 1358

matrix. In one instance, European Kestrels (Falco tinnunculus) were found to have greater 1359

relatedness in urban individuals as compared to rural individuals, despite similar allelic 1360

diversity in the two populations (Riegert et al. 2010) and genetic differentiation between 1361

urban and rural populations (Rutkowski et al. 2006). Within owls specifically, studies have 1362

not addressed the relative impacts of different matrix types but some have examined 1363

impacts of anthropogenic habitat fragmentation generally. Mediterranean Eagle Owls 1364

(Bubo bubo) have shown evidence of substantial population structure within a small 1365

geographic area of Spain as a consequence of anthropogenic habitat fragmentation (León-1366

Ortega et al., 2014). Additionally, closely related individuals have been found in mated pairs 1367

of Powerful Owls in urban fringe areas but not in adjacent intact woodlands (Hogan and 1368

Cooke, 2010). Reductions in genetic diversity and connectivity in susceptible taxa like 1369

predatory birds may serve as an early indicator of ecosystem decay in fragmented 1370

landscapes. Further investigation of these factors has the potential to identify landscapes at 1371

risk of reductions in biodiversity at higher trophic levels as a consequence of extinction debt 1372

incurred via habitat fragmentation. 1373

1374

Declines in Australian Boobook Abundance 1375

The Australian Boobook (Ninox boobook) is a common and widespread owl species 1376

found across most of continental Australia but apparent range-wide declines have 1377

prompted calls to investigate potential drivers of reductions in abundance (BirdLife 1378

Australia, 2015). Consistent trends in historical accounts of boobook abundance support 1379

74

the hypothesis that they have specifically declined in abundance in the Perth metropolitan 1380

area in Western Australia. Alexander (1921) described boobooks as resident in Perth and 1381

referred to them as “the common owl of the district.” Serventy (1948) repeated Alexander’s 1382

assessment but with an added qualifier: “the common owl of the district, but not frequently 1383

heard in the immediate vicinity of Perth.” Several decades later, Storr & Johnstone (1988) 1384

described the boobook as a moderately common passage migrant in Perth with no breeding 1385

records but “possibly also an uncommon resident.” Most recently, Stranger (2003) directly 1386

suggested a decline in boobook abundance in urban areas of the Swan Coastal Plain, stating 1387

that they “formerly ranged broadly over the plain, but [are] now rarer in the suburbs.” 1388

While these statements are only qualitative, they paint a picture of a population in decline 1389

in conjunction with increased urbanization. A similar account from agricultural landscapes 1390

inland of Perth suggests a reduction in boobook abundance in the Shire of Northam 1391

coinciding with extensive clearing of bushland for agriculture in the 1930s: “Uncommon, 1392

widespread in small numbers but not heard as often as during 1930s” (Masters and 1393

Milhinch, 1974). A more quantitative study has also demonstrated a negative correlation 1394

between boobook abundance and intensity of urban development (Weaving et al., 2011). 1395

While secondary exposure to anticoagulant rodenticides likely explains some of the 1396

decrease in observations in urban and peri-urban environments since the 1980s (Lohr, 1397

2018), the drivers of this pattern in other areas of the country are not clear and exploration 1398

of potential genetic impacts of fragmentation is warranted. While all of these observations 1399

indicate declines related to conversion of natural landscapes to human land uses, it is 1400

unclear whether these responses are related to fragmentation or simply the loss of usable 1401

space. 1402

1403

Boobook Movement and Responses to Fragmentation 1404

At present, a variety of conflicting views on boobook movement and dispersal 1405

patterns exist. Most sources suggest they are year-round residents where they occur 1406

(Saunders & Ingram 1995; Higgins 1999; König et al. 1999), especially in cooler temperate 1407

areas (Olsen & Taylor 2001; Olsen et al. 2011). Within the southwest of Western Australia, 1408

Storr and Johnstone (1988) describe the boobook as a “passage migrant” on the Swan 1409

Coastal Plain, though other sources suggest that perceived migrations may merely reflect a 1410

75

decrease in detectability due to a reduction in calling during the non-breeding season (Olsen 1411

and Debus, 2013) or short-distance home range shifts by some females during the non-1412

breeding season (Olsen and Taylor, 2001). McDonald and Pavey (2014) estimated 1413

movements of ≥ 32 km by boobooks in response to an arid zone rodent plague, potentially 1414

demonstrating longer-range temporary shifts in home range than suggested by Olsen and 1415

Taylor (2001). This estimate was based on the distance between their observations and the 1416

nearest assumed breeding habitat (Mcdonald and Pavey, 2014). These movements would 1417

have occurred across a landscape composed entirely of native bushland and may not be 1418

indicative of boobook movement patterns across fragmented landscapes. In summary, the 1419

existing evidence relating to boobook dispersal capacity across fragmented environments is 1420

limited and inconclusive. 1421

Across their range, boobooks are subject to predation by larger raptors, including 1422

Wedge-tailed Eagles (Aquila audax) (Cherriman, 2007) Grey Goshawks (Accipiter 1423

novaehollandiae) (Olsen et al., 1990), Brown Goshawks (Accipiter fasciatus) (Czechura et al. 1424

1987), and larger owls (Debus, 2009). Therefore, we hypothesised that they would be less 1425

likely to cross large sparsely-vegetated agricultural areas where they could be exposed to 1426

greater predation risk. That is, the matrix could be considered more hostile and less 1427

permeable in agricultural regions than in urban areas. Under these circumstances, 1428

fragmentation by agriculture in the wheatbelt could be functionally different to urban 1429

fragmentation in Perth with regard to dispersal and subsequent genetic impacts. 1430

Determination of genetic connectivity in boobooks via genetic analysis of individuals 1431

on a landscape scale will help settle long-standing speculation about the basic biology of this 1432

species and inform the management of an ecologically important and widespread avian 1433

carnivore. We aimed to determine whether potential differences in permeability of 1434

different types of anthropogenically-altered landscapes impacted genetic diversity and gene 1435

flow in a common but declining predatory bird by examining patterns of spatial genetic 1436

structure and corroborating these data with movement data derived from mark-recapture 1437

studies. We predicted that barriers to gene flow would occur in both urban and agricultural 1438

landscapes but would be more apparent in habitat fragmented by agricultural land use due 1439

to reduced dispersal capacity across a more hostile matrix. 1440

76

Methods 1441

Juvenile Dispersal 1442

To directly assess boobook dispersal capacity across fragmented habitats, we captured 1443

boobooks as nestlings or recent fledglings within their natal territory. Each young boobook 1444

was fitted with an individually-numbered stainless steel band issued by the Australian Bird 1445

and Bat Banding Scheme (ABBBS) to allow subsequent identification if re-sighted alive or 1446

recovered dead. A total of 17 boobooks from seven family groups were captured and 1447

banded. Of these, five individuals were re-sighted or found dead elsewhere. Location data 1448

submitted to the ABBBS by members of the public were then used to estimate dispersal 1449

distances. We also accessed data from the ABBBS from other banding studies elsewhere in 1450

Australia. We only included records of healthy birds captured in the wild to avoid potential 1451

bias from records of rehabilitated birds which may have behaved abnormally or been 1452

released away from the location where they were found. We found only eight additional 1453

qualifying instances of boobooks in Australia being banded as juveniles or nestlings and 1454

subsequently being resighted. One of these records was removed because the boobook 1455

was later recovered dead and still in the nest, leaving 12 available records, including those 1456

generated by our study. 1457

Genetic Sample Collection 1458

Western Australia is the largest state in Australia and covers an area of 1459

approximately 2,529,875 km² and makes up roughly the western third of the continent of 1460

Australia. We opportunistically collected blood and tissue samples from across the entirety 1461

of the state (Figure 4.1). We attempted to focus collection effort on three areas: the Perth 1462

metropolitan area, adjacent areas of continuous bushland in the Perth Hills, and agricultural 1463

areas in the agricultural Wheatbelt region in the vicinity the town of Kellerberrin 1464

approximately 200km east of Perth, in order to examine genetic structure across three 1465

distinct habitat types. 1466

77

1467

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

We used several methods to collect genetic information. Live Australian Boobooks 1473

were captured using a noose pole (Olsen et al. 2011) at night in conjunction with audio lures 1474

while conducting occupancy surveys across landscapes dominated by urban, bushland, and 1475

agricultural habitats. Additional wild boobooks were captured opportunistically using a 1476

noose pole when roosting individuals and family groups were reported by volunteers during 1477

the day. Blood was also collected from live boobooks held by wildlife rehabilitators along 1478

with information about where the boobook was originally found. Blood was drawn from the 1479

right jugular vein of each captured boobook using an insulin syringe with a 25G needle 1480

designed for subcutaneous use. In larger birds where more than a single capillary tube of 1481

blood is required, it is preferable to take blood from the right jugular vein, as this reduces 1482

handling time and risk of hematoma relative to sampling from the brachial vein (Owen, 1483

2011). The blood was refrigerated and allowed to coagulate for at least 24 hours prior to 1484

78

being centrifuged at 13000 RPM for 10 minutes. The resulting serum was removed for 1485

disease testing and the remaining material was frozen at -20°C for later genetic analysis. 1486

Additional samples were taken from boobook carcasses and shed feathers solicited 1487

from private citizens through BirdLife WA and a network of volunteers. Feathers were 1488

stored in paper envelopes at -20°C. All carcasses were stored frozen at -20°C until 1489

dissection when samples of muscle tissue were removed and stored in 100% ethanol for 1490

later analysis. 1491

Boobooks (n=137) were placed into one of six predefined regions based on 1492

similarities in geography and landscape type. The category ‘Exurbs’ (n=28) included 1493

individuals collected in areas immediately surrounding but not within the Perth 1494

Metropolitan area. ‘Perth Hills’ specimens (n=8) originated in an area of continuous forest 1495

east of Perth. Birds placed in the ‘Perth Metro’ category (n=71) originated in urban and 1496

suburban areas of Perth. Some boobooks were obtained from the Goldfields and Pilbara 1497

regions of Western Australia and were placed together in the ‘Remote WA’ (n=4) category. 1498

Boobooks from wetter, cooler, forested climates to the south of Perth were placed in the 1499

‘Southwest WA’ (n=17) category. The ‘Wheatbelt’ (n=9) category included all boobooks 1500

from highly-fragmented agricultural landscapes in the WA wheatbelt. 1501

1502

Genetic Analysis 1503

Microsatellites are commonly used in population genetic studies, particularly in bird 1504

species (Moura et al., 2017) for the purpose of individual fingerprinting, determining 1505

parentage, and exploring genetic variation and its spatial structure (Guichoux et al. 2011). 1506

Twenty microsatellite loci have been described for the Powerful Owl (Ninox strenua) and 19 1507

of these markers have been shown to be polymorphic in Australian Boobooks (Hogan et al. 1508

2009). Hogan et al. (2009) suggested these markers would be useful in genetic studies of all 1509

Ninox species tested. We used a subset of 15 loci microsatellites developed by Hogan et al. 1510

(2009) and after optimisation, nine (Nst02, Nst08, Nst11, Nst13, Nst14, Nst15, Nst16, Nst18, 1511

and Nst19; Table 4.1) were used to examine whether connectivity differed between the two 1512

types of anthropogenic landscapes. 1513

79

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

Locus GenBank Accession

no.

Primer (5`→3`) Reverse Core repetitive unit

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

Australian Boobook DNA was extracted from either a 2mm2 piece of muscle tissue or 1517

50µl of blood using a salting out technique described by (Miller et al., 1988) and re-1518

suspended in 100µl of amplification grade water. 1519

For amplification of microsatellite loci approximately 30 ng of genomic DNA was 1520

amplified by PCR with 5X polymerase buffer containing dNTPs (Fisher Biotec, Perth WA) 1521

either 1, 1.5 or 2 mM MgCl2 (Table 4.1), 0.2 µM F unlabelled primer and 0.4 µM of R primer 1522

and M13 labelled primer (Table 4.1), and 0.5 U Taq (Fisher Biotec) in a 10 µl reaction 1523

volume. PCR was performed on a Veriti thermocycler (Applied Biosystems). All samples 1524

were run on ABI 3500 Genetic Analyzer (Life Technologies) and scored using Genious V7.1 1525

(Biomatters, http://www.genious.com/). The following sets of loci were pooled together to 1526

run on the sequencer: (Nst02, Nst08, Nst11), (Nst13, Nst14, Nst15), (Nst16, Nst18, Nst19). 1527

Control samples were run in each PCR run to ensure compatibility between data used in the 1528

analysis. 1529

80

For loci Nst02 and Nst18, PCR conditions were an initial denaturation step at 94 °C 1530

for 5 minutes, followed by 30 cycles of denaturation at 94 °C for 30 seconds, annealing for 1531

45 seconds at 50 °C, and extension of 45 seconds at 72 °C with a final extension of 5 minutes 1532

at 72 °C. This was followed by another 8 cycles of denaturation at 94 °C for 30 seconds, 1533

annealing at 53 °C for 45 seconds and extension at 72 °C for 45 seconds. The last cycle was 1534

followed by final extension at 72 °C for 10 minutes. All other loci PCR conditions were an 1535

initial denaturation step at 94 °C for 5 minutes, followed by 4 touch down cycles of 1536

denaturation at 94 °C for 30 seconds, annealing for 45 seconds at 60-54 °C, and extension of 1537

45 seconds at 72 °C with a final extension of 5 minutes at 72 °C. This was followed by 1538

another 25 cycles of denaturation at 94 °C for 30 seconds, annealing at 54 °C for 45 seconds 1539

and extension at 72 °C for 45 seconds. The last cycle was followed by final extension at 72 1540

°C for 5 minutes. This was followed by another 8 cycles of denaturation at 94 °C for 30 1541

seconds, annealing at 53 °C for 45 seconds and extension at 72 °C for 45 seconds. The last 1542

cycle was followed by final extension at 72 °C for 10 minutes. 1543

Statistical Analysis 1544

Data for boobook owls were analysed at nine microsatellite loci described by Hogan 1545

et al. (2007) and Hogan et al. (2009). However, one locus was excluded from analysis 1546

(Nst14) due to a high frequency of genotyping failures, leaving eight microsatellite loci 1547

available for use in the results presented here. Highly related individuals, known offspring 1548

(sensu Wang 2018), and boobooks of unknown geographic origin were removed from 1549

analysis resulting in a sample size of 137 adult individuals (Appendix 4.1). We conducted 1550

Mantel tests and spatial autocorrelation using a subset of these individuals. Boobooks of 1551

known regional origin lacking precise location data were removed. Additionally, when more 1552

than one individual was sampled at a single location (usually in the case of mated pairs) one 1553

individual was randomly removed from analysis. This left 124 individuals available for the 1554

Mantel test. To examine spatial autocorrelation a subset of these individuals from the Perth 1555

Metro, Exurbs, and Perth Hills regions (n=98) were used because of the high density of 1556

sampling within these regions. 1557

To examine genetic relationships among groups of individuals and potential 1558

populations we conducted analysis of molecular variance (AMOVA) and principal 1559

coordinates analysis (PCoA) in GenAlEx6.502 (Peakall and Smouse, 2012, 2006). GenAlEx 1560

81

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

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

(HO), mean unbiased expected heterozygosity across all alleles (uHE), and fixation index (F). 1563

We also used GenAlEx to examine trends in isolation by distance using a Mantel test using 1564

all individuals with known coordinates and another Mantel test including only boobooks 1565

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

Jost’s DST, and Nm between regions. Spatial autocorrelation was tested in GenAlEx using 1567

even sample classes of n=200. We assessed genetic structuring using the program 1568

STRUCTURE 2.3.4 (Hubisz et al., 2009) using a burn-in of 100,000 steps and a MCMC of 1569

1,000,000 steps. We conducted 20 runs each assuming a different number of genetic 1570

clusters (K=1-6). We used CLUMPAK (Kopelman et al., 2015) to visually depict STRUCTURE 1571

outputs. STRUCTURE HARVESTER Web v0.6.94 (Earl and vonHoldt, 2012) was used to 1572

estimate the most probable number of genetic clusters using the Evanno et al. (2005) delta 1573

K method. 1574

Results 1575

Direct Measurement of Dispersal 1576

Across all 12 records, the average recorded distance between original capture 1577

location and subsequent observation in fledgling and nestling boobooks was approximately 1578

10.5km with a maximum recorded movement of 52 km (Table 4.2). In our study, juvenile 1579

boobooks were observed moving an average of 8km and up to 26 km from their capture 1580

site. All the captures and re-sightings of nestlings and fledglings from our study occurred 1581

within the Perth Metropolitan area across urban and suburban habitat.1582

82

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. 1583 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). 1584 Western Australian data from re-sightings and recoveries of boobooks captured as part of this study. 1585

Date Banded State/Territory Days elapsed between capture

and recovery

Distance travelled (kms) between capture and

recovery

Recovery Method

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

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

Source of variation Degrees of freedom

Sum of squares

Mean squares

Estimate of 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%

83

1589

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

1594

1595

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

-0.080

-0.060

-0.040

-0.020

0.000

0.020

0.040

0.060

0.080

0.100

3 7 9 11 13 15 17 19 21 23 25 27 29 31 33 35 38 42 46 53 65

r

Distance Class (End Point in kms)

r

U

L

84

Indirect Estimation of Dispersal 1600

All results suggested that there was no meaningful spatial genetic structuring in the 1601

population of boobooks we sampled. The Mantel tests did not detect a meaningful 1602

correlation between genetic and geographic distances in the entire group of boobooks 1603

sampled (Rxy=0.046, p=0.194) or within the metropolitan area (Rxy= 0.082, p=0.070). This 1604

result was corroborated by spatial autocorrelation analysis of boobooks from the Perth 1605

Metro, Exurbs, and Perth Hills regions which did not indicate significant genetic structure at 1606

any distance class (Figure 4.2). PCoA initially showed three distinct genetic clusters with no 1607

apparent correlation with hypothetical regions, and the first two axes explaining only 1608

15.91% of variance (Figure 4.3). Interrogation of the data set revealed that the three 1609

clusters were defined by homozygotes of two common alleles and their heterozygotes 1610

(Figure 4.3). When the locus Nst15 was removed from the analysis, no clusters were 1611

discernible (Figure 4.4) and the first two axes explained only 14.01% of the variance. The 1612

apparent clusters when the locus Nst15 was included appeared to be a consequence of a 1613

combination of low allelic diversity at the locus Nst15 and little genetic structure in the 1614

other seven loci. A lack of genetic structuring was also indicated by AMOVA which 1615

determined that 99% of the total molecular variance was partitioned within regions and 1616

only 1% among regions (Table 4.3). Fixation index values for all regions were within or below 1617

the range reported in populations of another small owl species which were not found to be 1618

impacted by a genetic bottleneck (Proudfoot et al., 2006) (Table 4.4). Pairwise Fst values 1619

between regions were low overall with the highest values between the Remote WA region 1620

and the other regions, consistent with largest geographic distance (Table 4.5). Similar 1621

patterns were evident in the estimated number of migrants per generation between regions 1622

(Table 4.5). However, even the highest values detected were still relatively low, particularly 1623

when taking into context the substantial geographic distance between the Remote WA 1624

collection locations and other regions and the large geographic area over which Remote WA 1625

specimens were obtained. Pairwise Jost’s DST values were also low between regions with 1626

the only significant value detected between the “Exurbs” and “Perth Metro” regions (Dst = 1627

0.027, P= 0.015) (Table 4.6). The statistical significance of this value is likely to be an 1628

artefact of the substantially larger samples sizes of these regions rather than indicative of a 1629

meaningful biological difference in the alleles present in the two regions. STRUCTURE 1630

results did not show any spatial genetic clustering (Figure 4.5). Low Delta K values also 1631

85

support a lack of spatial genetic structure (Figure 4.6). A single genetic cluster was 1632

supported by mean LnProb values obtained using CLUMPAK (Appendix 4.2) and STRUCTURE 1633

HARVESTER (Appendix 4.3). 1634

1635

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

Co

ord

. 2

Coord. 1

Principal Coordinates (PCoA)

Exurbs

Perth Hills

Perth Metro

Remote WA

Southwest WA

Wheatbelt

86

1642

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

1651

87

1652

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 1653 Australian Boobooks (Ninox boobook) sampled from Western Australia. 1654

1655

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

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

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

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

88

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

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

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

Discussion 1666

Both direct (banding) and indirect (genetic analysis) estimates of dispersal indicated 1667

widespread connectivity across all sampled populations despite extensive historical clearing 1668

of bushland in urban and agricultural landscapes. All statistical tests performed indicate a 1669

single admixed population of boobooks across all areas sampled. This result is consistent 1670

with a previous study which showed very little phylogenetic distinction between putative 1671

boobook subspecies across continental Australia (Gwee et al., 2017). The slightly higher Fst 1672

values observed between boobooks in the “Remote WA” group and other groups are likely a 1673

consequence of the group’s small sample size and the large geographic area from which the 1674

samples were derived. Alternately, weak isolation by distance across a large geographic 1675

area could explain this trend. 1676

The weak spatial genetic structuring both across Western Australia and within and 1677

between fragmented habitats is likely caused by effective movement between remnant 1678

habitat patches by dispersing juveniles. The genetic connectivity observed between 1679

89

fragmented landscapes and adjacent intact landscapes suggests historical movement 1680

between all areas despite extensive clearing over a long period of time (Saunders, 1989) 1681

while the observed capacity in our banding studies, of juvenile boobooks to disperse across 1682

substantial distances within fragmented urban landscapes, demonstrates that this type of 1683

habitat alteration does not constitute a barrier to juvenile dispersal. This result is consistent 1684

with dispersal patterns observed in other owl species. In a telemetry study of Burrowing 1685

Owls (Athene cunicularia), fledglings dispersed an average of 14.9 km (range 0.2 km - 53.1 1686

km) from their natal nest (Rosier et al., 2006). Similar dispersal patterns were observed in 1687

Spotted Owls (Strix occidentalis) (LaHaye et al., 2001). Within Australia, congeneric 1688

Powerful Owls (Ninox strenua) have been observed dispersing up to 18 km from their natal 1689

nest across “urban fringe habitat” (Hogan and Cooke, 2010). Dispersal by juvenile boobooks 1690

of distances substantially greater than those between patches of bushland habitat provides 1691

a plausible explanation for the lack of genetic structuring observed in the boobooks tested. 1692

While only movements within regions were observed in this study, the long distance 1693

contemporary dispersal observed within the Perth Metro region suggests the capacity for 1694

substantial post-breeding dispersal between regions. This result is consistent with the 1695

genetic estimate of migrants per generation among regions, suggesting considerable 1696

historical dispersal of juvenile boobooks (Table 4.5). 1697

Additionally, in the course of the study, boobooks were frequently observed and 1698

captured in urban areas outside of remnant bushlands. In some instances boobooks were 1699

observed successfully fledging young in areas where their home range would be expected to 1700

encompass no bushland whatsoever and be composed entirely of moderately dense 1701

suburban housing and light commercial development. If highly anthropogenically-altered 1702

habitats are able to support successful breeding attempts, these habitats likely constitute 1703

usable space despite their high degree of alteration and would not constitute a barrier to 1704

dispersal. Detection of moreporks (Ninox novaeseelandiae) at 80% of bushland patches in 1705

an urban area in New Zealand (Morgan and Styche, 2012) and documented use of highly 1706

developed suburban habitat by a female boobook during the non-breeding season (Olsen 1707

and Taylor, 2001) supports the hypothesis that these highly-altered habitat types do not 1708

provide a barrier to dispersal in boobooks. It is unclear to what degree the majority 1709

components of agricultural landscapes are “usable habitat” for boobooks but, on one 1710

90

occasion, in the course of this study, a boobook was observed hunting along a road >1km 1711

from any bushland, tree line, or patch of native vegetation, suggesting that boobooks 1712

actively utilize resources in habitats which we initially hypothesized to function as a hostile 1713

matrix between patches of usable habitat. In two Australian passerine species, natural 1714

history traits associated with tolerance of the “hostile matrix” in a fragmented landscape 1715

were demonstrated to correlate with spatial patterns of genetic diversity (Shanahan et al. 1716

2011). Boobooks are generalist predators capable of utilizing a wide variety of habitat types 1717

and are clearly capable of juvenile dispersal across urban development. Their capacity to 1718

use a wide variety of habitat types including highly anthropogenically-altered landscapes 1719

likely facilitates connectivity across ostensibly “fragmented” habitat. The lack of resistance 1720

observed in fragmented landscapes in our study of booboks probably protects them from 1721

the negative genetic impacts of fragmentation. Recent modelling of Mexican Spotted Owl 1722

(Strix occidentalis lucida) gene flow across fragmented habitats suggests that landscape 1723

resistance was an important predictor of genetic distance between populations for species 1724

with high dispersal capacity in highly fragmented landscapes (Wan et al., 2018). Owl species 1725

with more specialised habitat and dietary requirements including Blakiston’s Fish Owls 1726

(Bubo blakistoni) (Omote et al., 2015), Spotted Owls (Strix occidentalis) (Haig et al., 2001), 1727

and the more closely related Powerful Owl (Ninox strenua) (Hogan and Cooke, 2010) have 1728

shown genetic bottlenecks and potentially dangerous levels of inbreeding in response to 1729

habitat fragmentation. 1730

The lack of evidence for inbreeding or isolation as a consequence of habitat 1731

fragmentation does not necessarily imply that populations of boobooks in landscapes 1732

fragmented by urban and agricultural developments are demographically healthy or self-1733

sustaining. Weak spatial genetic structuring would likely also be observed in scenarios 1734

where fragmented habitats function as ecological sinks supported by healthy populations in 1735

adjacent intact habitats. This scenario is potentially even more likely in species with a high 1736

tolerance for altered habitats and substantial dispersal capacity. At least in urban areas, 1737

recent studies suggest that anthropogenic mortality from road strikes and secondary 1738

poisoning with anticoagulant rodenticides may pose significant threats to boobooks (Lohr, 1739

2018). Future work examining differences in life history parameters including adult and 1740

91

juvenile mortality across multiple habitat types would be useful in determining the relative 1741

utility of highly anthropogenically altered landscapes as boobook habitat. 1742

Genetic isolation and subsequent inbreeding could potentially become a problem for 1743

boobooks in urban and agricultural landscapes in the future despite their observed current 1744

dispersal capacity from banding studies if insufficient breeding hollows are retained at a 1745

landscape scale. Nest hollow availability is the key habitat requirement across the 1746

boobook’s range (Olsen and Taylor, 2001; Taylor and Canberra Ornitholgists Group, 1992) 1747

and urban fragments contain fewer hollow-bearing trees than intact forested areas (Harper 1748

et al. 2005). While nest hollow limitation does not currently appear to negatively impact 1749

boobooks in the Perth Metro area or WA wheatbelt (M. T. Lohr, unpublished data), 1750

continuing loss of nesting hollows through land clearing for additional development, 1751

inappropriate fire regimes, removal of nest trees for safety reasons, and urban infill could 1752

potentially reduce hollow availability in the future. In Powerful Owls (Ninox strenua), Hogan 1753

& Cooke (2010) detected instances of close inbreeding in two out of four pairs on the edge 1754

of urban areas near Melbourne despite a demonstrated capacity for dispersal up to 18km. 1755

Conversely, all three pairs nesting in continuous forested habitat were found to be 1756

unrelated (Hogan and Cooke, 2010). Hogan & Cooke (2010) speculated that this pattern 1757

could be explained by a lack of habitat for juveniles to disperse to, and subsequent 1758

clustering of related individuals, largely as a consequence of insufficient nest hollow 1759

availability. If patterns of boobook nest hollow availability ultimately approach those of 1760

Powerful Owls, this could lead to a reduction in genetic diversity and inbreeding depression 1761

over time in fragmented habitat types, even if boobooks are capable of dispersal between 1762

patches. However, if threatening processes and limiting factors in fragmented habitats are 1763

sufficiently addressed, both genetics and movement data suggest that boobooks should be 1764

capable of rapid recolonization and demographic recovery. 1765

Acknowledgments 1766

This project was supported financially by The Holsworth Wildlife Research 1767

Endowment via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird 1768

Research Award, and the Edith Cowan University School of Science Postgraduate Student 1769

Support Award. We thank Dr. Jamie Tedeschi for advice and technical assistance in 1770

laboratory work. We especially appreciate the contribution of boobook banding data by 1771

92

Jerry Olsen. Our research would not have been possible without contributions of samples 1772

and access to live birds provided by Kanyana Wildlife Rehabilitation, Native Animal Rescue, 1773

Native ARC, Nature Conservation Margaret River Region, Eagles Heritage Wildlife Centre, 1774

and many individual volunteers especially Simon Cherriman, Angela Febey, Amanda Payne, 1775

Stuart Payne, and Warren Goodwin.1776

93

Appendix 4.A A complete listing of the samples used in the analysis of microsatellite DNA polymorphisms, 1777

including the identification number (Individual ID), sample source, collection dates, collection locations 1778

(decimal lat/long), sampling locations/regions and age at sampling of Australian Boobooks used in this study. 1779

HY=hatch year, SY=second year, AHY=after hatch year, ASY=after second year. 1780

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

94

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

95

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

96

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

97

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

98

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

99

Appendix 4.B CLUMPAK results showing median values of the natural 1781

log of the probability of the number of genetic clusters (K=1-6) in 1782

Australian Boobooks sampled in Western Australia. 1783

1784

1785

Appendix 4.C STRUCTURE HARVESTER output indicating the highest 1786

probability for K=1 in boobooks sampled in Western Australia. 1787

1788

100

Chapter 5 Artificial nest box supplementation does not affect 1789

Australian boobook (Ninox boobook) occupancy in fragmented 1790

habitats in south-western Australia 1791

1792

Lohr, M. T., S. Cherriman, A. H. Burbidge, and R. A. Davis. Artificial nest box supplementation 1793

does not affect Australian boobook (Ninox boobook) occupancy in fragmented habitats 1794

in south-western Australia. Wildlife Research. (In Review). 1795

Abstract 1796

Nest hollows are critical elements of usable habitat for many wildlife species 1797

worldwide, particularly in Australia. Loss of hollows due to anthropogenic processes and 1798

competition with introduced species over remaining hollows are key threats to hollow-1799

nesting species in landscapes dominated by urban and agricultural development. 1800

Supplementation with artificial nest boxes has been suggested as a method to mitigate 1801

these threats but the efficacy of this technique has seldom been evaluated. The hollow-1802

nesting Australian Boobook (Ninox boobook), a small owl, has experienced a nearly range-1803

wide decline for reasons that are not well understood. We aimed to determine the utility of 1804

nest- box supplementation as a conservation action for boobooks and the influence of nest- 1805

box supplementation on potentially competing species across two different types of 1806

fragmented landscape. We monitored boobook occupancy in bushland fragments in urban 1807

and agricultural landscapes as well as in areas of continuous bushland before and after nest- 1808

box installation. Monitoring protocols involved nocturnal point counts and broadcast 1809

recordings of boobook calls and were based on methods used in previous owl surveys 1810

overlapping our study areas. We also used a pole- mounted video camera to record species 1811

using nest boxes during the boobook breeding season over the course of two years. Nest- 1812

box supplementation did not increase boobook occupancy at monitored sites over the 1813

period of this study, though one box was used successfully. Nest boxes were more 1814

frequently utilized by alien and overabundant native bird species. The ability of boobooks 1815

to use small hollows and possibly evict competing species probably insulates them from the 1816

impacts of hollow loss relative to other obligate hollow-nesting species. 1817

101

Introduction 1818

Tree hollows are used for shelter and nesting by taxonomically diverse wildlife 1819

species worldwide and are often a critical component of habitat for those species (Martin 1820

and Eadie 1999; Isaac et al. 2014). As a result, a shortage of available hollows can limit 1821

abundance of hollow-nesting species (Newton, 1994). The loss of nest hollows is an issue of 1822

particular conservation concern in Australia with 302 species of vertebrate recorded as 1823

using hollows for shelter or nesting (Gibbons et al. 2002). Eleven per cent of all bird species 1824

in Australia are classified as obligate hollow nesters compared to 5% in Europe, 4% in North 1825

America, and 6% in Africa (Newton, 1994). The absence of primary hollow-excavating 1826

vertebrate fauna like woodpeckers (Picidae) in Australia leaves hollow formation primarily 1827

dependent on stochastic processes (Saunders et al. 1982) including fire, decay by fungal or 1828

insect attack, and mechanical damage from other trees, wind, or lightning (Fox et al. 2009). 1829

Termites play a major role in facilitating rot in heartwood and subsequent excavation of 1830

hollows throughout Australian woodlands (Gibbons et al., 2000). Hollow formation can take 1831

more than 150 years in some trees (Harper et al. 2005) but may occur more quickly in other 1832

tree species (Whitford, 2002). Tree hollows are currently being lost in Australia faster than 1833

they are being replaced (Lindenmayer et al. 1997). The confluence of these factors makes 1834

nest hollow availability a critical and possibly limiting factor in the habitat requirements of 1835

many Australian wildlife species. 1836

Fragmentation of woodlands by human land uses can increase the rate at which 1837

hollows and hollow-bearing trees are lost through a variety of mechanisms. In urban 1838

remnant woodlands, edge effects involving increased wind exposure may substantially 1839

impact the abundance of tree hollows and may impact the entirety of the fragment 1840

depending on its size (Harper et al. 2005). In a survey of tree hollow occurrence in urban 1841

remnant woodlands in Melbourne, Australia, Harper et al. (2005) found no hollows in 12 of 1842

44 survey sites and 64% of remnants contained fewer than six hollow-bearing trees per 1843

hectare which is “well below that contained in areas of un-logged non-urban forest”. Urban 1844

remnant forests in Sydney were also found to have fewer hollow-bearing trees than 1845

continuous forest (Davis et al. 2014). The removal of large trees for timber and firewood as 1846

part of past management practices has also substantially decreased the number of hollow-1847

bearing trees in some urban remnants (Harper et al. 2005) and has directly impacted some 1848

102

threatened bird species such as the Swift Parrot (Lathamus discolor) (Webb et al., 2018). 1849

The exclusion of fire from urban forest fragments and removal of large trees due to safety 1850

concerns, may also play a role in reducing hollow formation and persistence (Harper et al. 1851

2005). Conversely, the inappropriate use of fire has been noted as a key driver of hollow 1852

loss (Stojanovic et al., 2016) and in agricultural regions has combined with other stressors 1853

such as intentional bulldozing of nest trees, and lone trees in paddocks being blown over as 1854

a consequence of greater exposure to wind (Saunders et al., 2014). 1855

Consistent long-term decline in abundance of large nest hollows used by endangered 1856

Carnaby’s Black-Cockatoos (Calyptorhynchus latirostris) has also been observed in remnant 1857

bushlands in agricultural landscapes in Western Australia (Saunders et al., 2014). The 1858

relative paucity of available hollows in landscapes which have been intensively altered by 1859

humans may be a limiting factor for wildlife which would otherwise be capable of using 1860

remnant bushlands and could be a factor contributing to overall declines in biodiversity. 1861

Nest Competition and Predation 1862

Even where nest hollow abundance is high, competition from introduced and 1863

overabundant native species can reduce nest hollow availability for obligate hollow-nesting 1864

wildlife. In North America, range-wide decline in three bluebird species (Sialia spp.) has 1865

been partially attributed to competition for nest hollows from European Starlings (Sturnus 1866

vulgaris) and House Sparrows (Passer domesticus) (Newton, 1994). In Australia, the 1867

introduction of hollow-nesting Common Mynas (Acridotheres tristis) was found to be 1868

correlated with declines in three native hollow-nesting bird species in the Canberra area 1869

(Grarock et al. 2012). Introduced European honeybees have been recorded as excluding a 1870

wide variety of native marsupial species from nest boxes in Australia (Beyer and Goldingay, 1871

2006). Galahs (Eolophus roseicapilla) and Western Corellas (Cacatua pastinator) are native 1872

to Western Australia but are overabundant in some areas and are believed to negatively 1873

impact endangered Carnaby’s Black-Cockatoos through competition for scarce nesting 1874

hollows (Johnstone et al., 2015; Saunders and Doley, 2017). Predation by the introduced 1875

Sugar Glider (Petaurus breviceps) in Tasmania, Australia is the key cause of the decline of 1876

the endangered hollow-nesting Swift parrot (Stojanovic et al., 2014). Understanding 1877

interactions between native and introduced hollow nesting species will be of increasing 1878

103

importance to conserving native biodiversity in areas where fragmentation simultaneously 1879

decreases hollow availability and facilitates growth of populations of introduced species. 1880

Impacts of Nest Boxes in Conservation 1881

As a response to hollow limitation, artificial nest hollow or “nest box” installation 1882

programs have been used for research aimed at understanding important life history traits 1883

of specific populations and have been used as an effective conservation measure to stabilize 1884

some declining populations (Lambrechts et al. 2012). These programs have been an 1885

important part of recovery efforts for hollow breeding birds worldwide. Routing of artificial 1886

hollows into living trees has been an integral part of successful efforts to increase 1887

abundance of the endangered Red-cockaded Woodpecker (Leuconotopicus borealis) in the 1888

southeastern United States (Walters, 1991). Widespread nest box provisioning efforts by 1889

private organizations have been widely attributed as a major factor in the recovery of three 1890

species of bluebirds in North America (Newton, 1994). Nest boxes have also been used to 1891

increase barn owl populations in Israel (Kan et al. 2013), Malaysia (Duckett and Karuppiah 1892

1990; Puan et al. 2012), and India (Parshad, 1999) as part of efforts to reduce crop damage 1893

by rodents. In Australia, construction of nest boxes is currently used successfully to mitigate 1894

losses of natural hollows for Carnaby’s Black-Cockatoos in the Western Australian 1895

agricultural zone (Johnstone et al. 2015) and Glossy Black-Cockatoos (Calyptorhynchus 1896

lathami) on Kangaroo Island (Mooney and Pedler, 2005) and has been used as a 1897

conservation tool in managing Critically Endangered Orange-bellied Parrots (Neophema 1898

chrysogaster) (Goldingay and Stevens, 2009) and Swift Parrots (Stojanovic et al., 2019). 1899

While most of these programs addressed lack of hollow availability, in some bird species, a 1900

variety of parameters impacting breeding success are higher in nest boxes than in natural 1901

hollows (Purcell et al. 1997). 1902

Nest boxes may not be a solution for all species, especially if nest hollow limitation is 1903

not the key cause of decline. For example, Loman (2006) found that nest hollow availability 1904

in small woodland patches was limiting for some obligate hollow-nesting passerine species 1905

but not others. In some instances, nest boxes may be preferred to natural nests and rapid 1906

adoption of nest boxes can give the appearance of nest limitation where there is none. For 1907

example, in one study, 83% of Tawny Owl pairs switched from natural nest sites to nest 1908

boxes within the year they were provided and 100% of pairs switched within four years but 1909

104

breeding density did not appear to change as a result of nest box provisioning (Petty, 1992). 1910

Purple Martins (Progne subis) in North America provide an even more extreme example of 1911

this dynamic. The eastern population of Purple Martins has used nesting structures 1912

provided by humans since prior to European colonization (Speck, 1941) and is now almost 1913

completely dependent on artificial nesting hollows constructed by humans (Morton et al. 1914

1990). While human-provisioned nest hollows clearly benefit this species, the potential 1915

risks of a population’s near-complete dependence on nest sites provided by humans are 1916

evident. In instances where nest boxes are preferred to natural hollows but are associated 1917

with lower nesting success they may even function as ecological traps (Klein et al. 2007, 1918

Heinshohn et al., 2015). Perhaps most fundamentally, nest box supplementation will not 1919

result in increases in abundance of the target species unless other resource requirements 1920

are already met (Durant et al. 2009). These factors should be considered before 1921

implementing or encouraging large-scale nest box programs and when evaluating the 1922

results of these programs. 1923

Knowledge Gaps 1924

Despite a large body of research on nest box impacts on native mammals and use of 1925

nest boxes in conservation efforts for cockatoos, few studies have focused on use of nest 1926

boxes by predatory birds in Australia. In a review of literature regarding nest box use by 1927

Australian bird species, only one of 17 species listed as having been studied was a predatory 1928

bird (Goldingay and Stevens, 2009). This study was conducted on a small hybrid population 1929

of boobooks on Norfolk Island and was an overview of conservation efforts rather than an 1930

empirical study of nest box impacts (Olsen, 1996). Another major knowledge gap relating to 1931

nest box impacts involves their use in developed areas. Less than 5% of Australian studies 1932

on the use of natural and artificial hollows have been conducted in urban landscapes 1933

(Durant, 2006). 1934

Despite the lack of studies relating to use of nest boxes by urban birds generally and 1935

Australian raptors specifically, artificial nest hollows have already been promoted as a 1936

conservation measure for urban raptors. Provision of nest boxes was suggested to improve 1937

Powerful Owl habitat in urban environments where scarcity of suitable nest hollows may be 1938

limiting abundance (Isaac et al. 2008). In one instance, subsequent localized nest box 1939

placement resulted in successful breeding of a nesting pair (McNabb and Keating, 2008; 1940

105

McNabb and Greenwood, 2011). While this particular effort was well justified and the 1941

result of this action is encouraging, the unregulated implementation of untested 1942

conservation actions intended to benefit sensitive species is concerning. Nest boxes 1943

intended for use by a wide variety of wildlife species are already commercially available 1944

from local businesses and instructions and plans are readily available online and are actively 1945

promoted for owls by the WA government Department of Biodiversity, Conservation and 1946

Attractions (Hussey, 1997). Both options are promoted as broadly beneficial to native 1947

wildlife despite a lack of rigorous testing for most species. Klein et al. (2007) suggested that 1948

correlation between increased breeding abundance and nest box provisioning should be 1949

proven prior to use of nest boxes as a conservation strategy. The widespread promotion 1950

and use of nest boxes necessitates studies addressing impacts of nest boxes on bird 1951

populations broadly and particularly on predatory birds and birds using urban areas. 1952

We studied the small owl, the Australian boobook in south-western Australia, as a 1953

model to examine whether nest box provisioning can increase occupancy by this species in 1954

human-altered landscapes. Australian boobook’s are an ideal study species as they are 1955

widespread but a 2015 report on population trends in Australian birds identified a serious 1956

decline in Australian boobook numbers from 1999-2013 and recommended that “further 1957

investigation is needed to understand the factors that are driving this consistent decline 1958

across regions” (BirdLife Australia, 2015). Nest hollow availability is believed to be the key 1959

habitat requirement across the boobook’s range (Olsen and Taylor, 2001; Taylor and 1960

Canberra Ornitholgists Group, 1992) and loss of tree hollows has been cited as one of the 1961

reasons for its decline in some areas (Debus, 2009). In the single published study involving 1962

nest box use by boobooks, lack of nesting hollows was implicated as one of the major 1963

factors contributing to the near extinction of Norfolk Island boobooks (Ninox 1964

novaseelandiae undulata) and nest boxes were a key tool used in its recovery program 1965

(Olsen, 1996). Boobook occurrence has been observed to correlate negatively with 1966

increased density of sealed roads and positively with forest cover, and nest hollow 1967

availability has been hypothesized as the factor driving differences in boobook abundance 1968

between urban and forested landscapes in and around Melbourne, Australia (Weaving et 1969

al., 2011). Urban fragments generally contain fewer hollow bearing trees than intact 1970

forested areas (Harper et al. 2005). Likewise, in the agricultural “wheatbelt” of Western 1971

106

Australia, loss of nest hollows is said to be one of the most important challenges facing 1972

wildlife conservation (Johnstone et al. 2015). Examination of patterns of nest box use by 1973

boobooks and its relationship with site occupancy across these two habitat types is 1974

necessary to understand their potential utility in the conservation of this species. 1975

Specifically we aimed to investigate whether Australian boobook occupancy in 1976

fragmented landscape types (agricultural and urban) was altered by providing nest boxes. 1977

Our hypothesis was that nest hollows would be limiting in agricultural and urban landscapes 1978

and that nest boxes would be quickly taken up by Australian boobooks. 1979

1980

Methods 1981

Study Sites 1982

To determine the impacts of nest box installation on site occupancy, surveys were 1983

conducted in 2015 at >30 sites each in each of three categories of land use: urban remnant 1984

bushlands, agricultural remnant bushlands, and areas of continuous bushland . Sites were 1985

located along the same approximate latitude in an area of south western Western Australia 1986

with a Mediterranean climate (Figure 5.1). Urban sites (n=35) found across the Perth 1987

metropolitan area were composed of bushland reserves managed by city governments or 1988

the Botanic Parks and Gardens Authority. Most sites were open woodlands dominated by 1989

Banksia sp., Eucalyptus gomphocephala, or E. rudis. Agricultural sites (n= 33) included both 1990

privately-owned bushlands and sites managed by the Western Australian Department of 1991

Biodiversity, Conservation and Attractions. All were within approximately 60km of the town 1992

of Kellerberrin, Western Australia. Dominant vegetation across these sites included Acacia 1993

acuminata, Eucalyptus capillosa, E. loxophleba, and E. salmonophloia. Continuous bushland 1994

sites (n=34) were located between the Perth Metropolitan area and areas of extensive 1995

agricultural development. They were bounded by the Great Eastern and Great Southern 1996

Highways to the North and the Brookton Highway to the South. Dominant vegetation in 1997

these sites was primarily Eucalyptus wandoo, E. marginata, and Corymbia calophylla. Intact 1998

bushland sites were included in surveys as a baseline against which to compare the efficacy 1999

of nest box supplementation as a management action intended to increase site occupancy. 2000

107

2001

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

Surveys 2004

In urban and agricultural bushlands, surveys were conducted 100m from a road or 2005

near the middle of the reserve in smaller reserves to reduce the impact of traffic noise on 2006

surveys. In continuous bushland areas, survey points were located approximately 5km apart 2007

to ensure independence. Baseline occupancy surveys were conducted in 2015 from 2008

September to December during the breeding season when boobooks call most frequently 2009

and are most easily detected (Olsen, 2011b). To maintain consistent detectability of 2010

boobooks, surveys were only conducted in the absence of rain and when estimations of 2011

wind speed were below a score of 3 on the Beaufort scale. Surveys consisted of passively 2012

listening for boobook vocalizations from a fixed point for 15 minutes followed by five 2013

minutes of intermittent broadcast of recorded boobook vocalizations in accordance with 2014

methodology used by Liddelow et al. (2002) to survey nocturnal birds in south-western WA. 2015

Immediately following the survey, the area was scanned using a 1000 lumen LED headlamp 2016

to detect any boobooks that had been attracted by the calls but had not vocalised. All sites 2017

108

were classified as “occupied” or “not occupied.” A subsequent round of surveys was 2018

conducted at all sites in September-December of 2016 after the installation of nest boxes to 2019

determine occupancy using the same methodology used during the previous breeding 2020

season. 2021

Nest box construction and placement 2022

Nest boxes were constructed using recycled, 18mm form-ply, a waterproof and long-2023

lasting material used mainly for concrete construction work. Each box consisted of a 2024

wooden cube measuring 300mm long and 300mm wide, and a depth ranging from 450mm 2025

at the front to 500mm at the rear, creating a forward-sloping roof. The dimensions of the 2026

box were chosen to reflect dimensions of active boobook nests observed by the authors and 2027

to deter Galahs, which prefer deeper boxes and are aggressive competitors for nest hollows. 2028

A hollow log-round of diameter 120-200mm was attached to the front of each box using 2029

screws fastened from the inside. This served to create a ‘verandah’ designed to protect the 2030

internal nest-chamber from weather, and also to prevent non-target species with 'heavy-2031

chewing' behaviour (e.g. Galahs) from enlarging the box entrance hole and potentially 2032

destroying it. A wooden lid with a c. 50mm overhang was attached with a hinge fitted to the 2033

rear, and the sides were reinforced with aluminium flashing, again to prevent chewing 2034

species from destroying the lid. Boxes were assembled in such a way to leave ’air slots’ 2035

~15mm wide beneath the lid on both sides, designed to facilitate air-flow and subsequent 2036

internal temperature fluctuation to deter feral honey bees, which have specific hive 2037

temperature requirements of 32-35˚C, from taking up residence. Two coats of pale-green, 2038

water-based exterior paint were applied to all external surfaces, to protect the sawn 2039

wooden edges from the elements and thus defer deterioration, and to help the boxes blend 2040

in with the natural environment. A layer of coarse woodchips c. 150mm deep was added to 2041

the inside of each box to create an internal nest chamber consisting of well-drained 2042

substrate that allows hollow-nesting species to scrape a shallow bowl in which eggs are 2043

deposited. Wood chips consisted of c. 20mm diameter pieces and were collected near 2044

installation sites. 2045

Nest boxes were installed in trees with multi-strand, galvanised wire (‘clothesline 2046

wire’) c. 4mm thick, threaded through plastic/rubber pipe (‘hosepipe’) to protect the tree's 2047

bark from wire damage. The installation process was carried out using the following steps: 2048

109

1) a small loop was created at one end of the wire, and the tail end threaded into one of two 2049

c. 8mm holes pre-drilled on the rear surface of the box, at each of its top corners; 2) the tail 2050

end was then threaded out through the second hole and the wire pulled through until it was 2051

tight; 3) after being threaded through a length of hosepipe, the main length of wire was 2052

looped horizontally around a solid, vertical section of trunk, being passed above an oblique 2053

or horizontal limb used to ‘hang’ the box and prevent it sliding down (Figure 5.2); the tail 2054

end was then threaded through the small loop at the back of the box and twitched into 2055

place for secure attachment. Sufficient length of wire was used so each box was ’strung’ 2056

firmly but not hung in such a way that left wire tightly constricting on the trunk. This 2057

method is similar to the ‘habisure system’ described in Franks and Franks (2006), and it 2058

ensures secondary (horizontal) growth of the tree’s trunk (i.e. limb thickening) can take 2059

place naturally. Permanent attachment methods involving fixings such as coach bolts or 2060

screws were avoided to 1) minimise injury to the tree’s vascular cambium that may lead to 2061

unnecessary infection or damage, and 2) ensure boxes were not ‘pushed off' as the tree 2062

trunk expands during secondary growth, resulting in the potential collapse of an occupied 2063

nest-site and/or a safety risk to passers-by. 2064

110

2065

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

Nest boxes were placed in fifteen sites in both urban and agricultural remnant 2067

bushlands which did not have boobook detections in the previous round of surveys. Nest 2068

boxes were installed in February 2016 shortly after the termination of the breeding season 2069

to allow adequate time for detection by boobooks prior to the following breeding season. 2070

All nest boxes were placed in the nearest suitable tree to the survey point in all 30 2071

experimental sites. All nest boxes were hung at a height below 11m to facilitate observation 2072

of their contents and greater than 4m because most published records indicate minimum 2073

nest heights above 3m for boobooks (Higgins, 1999) (Figure 5.3). 2074

111

2075

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

Nest Box Monitoring 2077

We examined the contents of all nest boxes for evidence of use by boobooks or 2078

potentially competing species. Nest box contents were viewed using a video camera 2079

(MiGear ExtremeX Sports Action Camera) mounted on an 8m telescoping fiberglass pole to 2080

record video footage of the inside of each nest box. All videos were viewed at the nest site 2081

112

to ensure that adequate footage of the nest boxes’ interior was obtained to allow 2082

identification of contents. Videos were retained for later review. Nest boxes were checked 2083

on three occasions during the breeding season in 2016 (July 24-26, October 7-9, and 2084

November 18-25) and once during the 2017 breeding season (September 27-29). In the 2085

2017 surveys, a single nest box at one of the urban sites was unavailable to be checked as it 2086

had been destroyed by a bushfire. 2087

Statistical Analysis 2088

We compared differences in territory occupancy in 2015 across all three habitat 2089

types prior to treatment using a Chi-square test with a post hoc pairwise test of 2090

independence for nominal data. We used McNemar's Chi-squared tests with continuity 2091

correction to examine differences in occupancy between years at treated and untreated 2092

sites. All tests were performed using RStudio 1.1.383 (RStudio, Inc., Boston, MA, USA). 2093

Results 2094

Prior to nest box treatment, boobooks were more commonly detected in continuous 2095

bushland sites (85.3%, n = 34) than in remnant bushlands in urban (30.8% n = 39) and 2096

agricultural (21.2%, n = 33) landscapes. Occupancy rates were significantly greater at 2097

continuous bushland sites than urban sites (p<0.001) or wheatbelt sites (p<0.001) but did 2098

not differ between urban and wheatbelt sites (p=0.517). No significant differences in 2099

occupancy were detected between years in any of the treated or untreated groups across all 2100

three habitat types (Table 5.1). However, non-significant increases in occupancy occurred in 2101

sites provided with nest boxes in both fragmented habitats while non-significant declines 2102

occurred in control sites in both urban and agricultural habitats (Table 5.1).2103

113

2104

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 2105 landscapes in Western Australia. 2106

Total sites No. Occupied

2015 % Occupied

2015 No. Occupied

2016 % Occupied

2016 McNemar's 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

114

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

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

Nest boxes were used by a total of six species (Table 5.2). The most commonly 2111

detected species utilizing nest boxes was the Australian Wood Duck (Chenonetta jubata). 2112

The only exotic species observed nesting in the nest boxes was the Laughing Kookaburra 2113

(Dacelo novaeguineae; introduced to Western Australian in the early 1900s). All nest boxes 2114

used by this species were in urban bushlands. Boobooks used one nest box, located in one 2115

of the urban bushlands, during the 2016 breeding season. Three large and healthy owlets 2116

were observed in this nest box. We assumed this nest box to have been successful because 2117

the age of the nestlings calculated using the equation given by Olsen et al. (2015) was 2118

greater than the average age at which boobooks fledge. 2119

Discussion 2120

Surveys 2121

The detection of boobooks in 85.3% of continuous bushland areas in both years is 2122

higher than in previous surveys conducted in similar forested areas of Western Australia 2123

during spring (Liddelow et al., 2002), in which boobooks were detected at only 61.5% of 2124

sites. Our use of boobook calls, which were not used in previous studies, likely improved 2125

our ability to detect boobooks that were present. It is also possible that broadcasting the 2126

calls of Barking Owls and Masked Owls in previous studies may have supressed boobook 2127

calling, as these species are larger and may compete with or prey on boobooks. Boobooks 2128

will sometimes stop calling in response to broadcast calls of Powerful Owls or Masked Owls 2129

(Debus, 2009). A study of boobooks in suburban and forested habitats around Melbourne, 2130

115

Victoria found higher boobook occupancy in forested sites (94%) than our study (85.3%) but 2131

substantially lower occupancy in suburban sites (13%) than was observed in our study prior 2132

to nest box supplementation (30.8%) (Weaving et al., 2011). The higher occupancy 2133

recorded by Weaving et al. (2011) in forested areas may simply be an artefact of greater 2134

sampling effort and the use of transects rather than point counts. However, the difference 2135

in occupancy rates in urban areas runs counter to what would be expected given the 2136

differences in methodologies, suggesting an actual difference. Some of the urban areas of 2137

Perth where our study was conducted have been developed more recently than the sites in 2138

Melbourne. This may mean that extinction debt generated as a result of fragmentation at 2139

urban sites in Perth has not been fully paid, which could explain the disparity in occupancy 2140

between the two studies. This hypothesis is consistent with the observation that despite 2141

nest box supplementation at some sites, boobook occupancy across all urban sites dropped 2142

from 30.8% to 20.5% between 2015 and 2016. If this hypothesis is correct and our 2143

occupancy estimates are low relative to those of Weaving et al. (2011) due to our fewer 2144

surveys and different methodology, further reductions in boobook abundance can be 2145

expected in the Perth Metropolitan area. 2146

While the decline observed in boobook occupancy at urban sites is difficult to 2147

substantiate due to low sample size and an insufficient number and duration of surveys at 2148

the same sites, it is consistent with national trends indicating a continental scale decline in 2149

boobook abundance (BirdLife Australia, 2015) and reductions in boobook occupancy in 2150

urban bushlands in Canberra (Olsen and Trost, 2015). Conversely, occupancy was roughly 2151

stable bushland fragments in agricultural landscapes and continuous bushland. Recent 2152

research on boobooks in Western Australia documented pervasive and sometimes lethal 2153

exposure to anticoagulant rodenticides associated with proximity to developed habitat, but 2154

not agricultural or bushland habitat (Lohr, 2018). Secondary anticoagulant poisoning is a 2155

plausible mechanism explaining the observed differences in population trajectories across 2156

these three landscape types. 2157

Nest Box Use 2158

The use of nest boxes primarily by introduced species and overabundant native 2159

species must be considered when evaluating the use of nest boxes as a tool in conservation. 2160

Laughing Kookaburras are not native to Western Australia and anecdotal accounts of 2161

116

negative impacts on breeding native passerines (Serventy, 1980) suggest that facilitating 2162

breeding of Laughing Kookaburras should not be encouraged. While Laughing Kookaburras 2163

have not been documented directly competing with boobooks for nest hollows, they have 2164

been observed directly killing roosting boobooks during the day on several occasions 2165

(Higgins, 1999). 2166

Aside from boobooks, all native species documented using nest boxes in this study 2167

are subject to control in some areas. Western Corellas and Galahs have been culled – 2168

sometimes in large numbers – as part of conservation efforts to reduce nest competition 2169

with endangered Carnaby’s Black-Cockatoos (Saunders and Doley, 2017). The use of nest 2170

boxes by abundant species may not be a desirable outcome in all circumstances. There is an 2171

open season on Australian Ringnecks (Barnardius zonarius) and corellas (Cacatua spp.) 2172

across most of south-western Australia and damage permits may be issued for Australian 2173

Wood Ducks (Chenonetta jubata) in agricultural areas (Department of Biodiversity, 2174

Conservation and Attractions 2019). Any nest box programs initiated to benefit a specific 2175

species should incorporate monitoring regimes and protocols for managing use by species 2176

which managers do not wish to facilitate. 2177

Several hypotheses potentially explain the minimal use of nest boxes by boobooks. 2178

Nest boxes were deliberately placed at unoccupied sites, so it is possible that boobooks 2179

were simply absent from these areas. However, in light of the substantial dispersal capacity 2180

of boobooks in our study areas inferred from genetic and banding data presented in Chapter 2181

4, this seems does not appear to be a likely explanation. It also seems unlikely that 2182

boobooks failed to use nest boxes due to an insufficient amount of time to locate the boxes. 2183

One box was located and utilized within the first year after installation but no boxes appear 2184

to have been used by boobooks in the following breeding season. Low abundance in 2185

fragmented habitats driven by factors other than nest site availability could also potentially 2186

explain low uptake of nest boxes. Alternately, it is possible that the box design is simply not 2187

favoured by boobooks. Some preference that is not currently understood could be at work. 2188

Subtle aspects of nest box construction can impact nest box use (Lambrechts et al., 2012). 2189

In some instances these preferences can be strong and unexpected. For example, in 2190

American Kestrels (Falco sparverius), nest box dimensions had a strong effect on uptake 2191

(Bortolotti, 1994). 2192

117

While it is possible that the nest box design may not have been ideal for boobooks, 2193

Australian boobooks, as a species, appear to be fairly plastic in their use of nest sites. They 2194

have been recorded using caves in treeless areas (Higgins, 1999) and old corvid nests 2195

(Sedgwick and Morrison, 1948) where tree hollows are not available. Boobooks are already 2196

known to use nest boxes and have been observed doing so in urban areas. Hogg and Skegg 2197

(1961) describe the successful nesting of a pair of New Zealand Moreporks – a closely 2198

related species – in a nest box adjacent to a noisy rifle range in Auckland, New Zealand. 2199

Prior to our study, two cases of boobooks using similarly-constructed nest boxes were 2200

documented specifically within the Perth metropolitan area. In one instance, boobooks 2201

successfully raised chicks (with artificial supplementation of food) in a nest box in the Perth 2202

suburb of Victoria Park (Wells, 2007). Beckingham (2012) also reported an adult boobook 2203

seen at the entrance of an artificial nest box placed in bushland near Lake Claremont and 2204

included a photo of a downy juvenile boobook observed nearby several weeks later. 2205

Another explanation for the low rate of nest box use by boobooks is that boobooks 2206

are not limited by nest hollow availability in either urban or agricultural bushland fragments. 2207

Despite assertion in previous literature that Australian boobooks are insectivorous, they are 2208

capable of preying on relatively large birds and mammals (Olsen, 2011a). As a consequence 2209

they are probably able to compete successfully with most other species likely to use a 2210

hollow of appropriate size. In the one occupied nest box, when chicks were temporarily 2211

removed for measurement, banding, and blood sampling, we observed the remains of 2212

Rainbow Lorikeets suggesting that the lorikeets are not likely to be effective in competing 2213

with boobooks for potential nest hollows. A similar instance was reported in which Galah 2214

nestlings were presumed to have been eaten by a boobook prior to the boobook nesting in 2215

their hollow. The remains of a male Australian Ringneck – another large hollow-nesting 2216

parrot – were subsequently found in the same nest with two boobook nestlings (Mack, 2217

1965). Other instances of boobooks taking over hollows actively used by Galahs have been 2218

reported (Schulze, 1966). Conversely, the authors have observed an instance of boobooks 2219

being evicted from a nest hollow following repeated harassment by Galahs. Common 2220

Mynas – another potential nest competitor – have also been observed harassing boobooks 2221

as they left their nest hollow but the same pair of boobooks was later photographed eating 2222

Common Mynas (Trost and Olsen, 2016). The closely related New Zealand Morepork has 2223

118

also been observed successfully evicting European Starlings from a nest hollow (Hogg and 2224

Skegg, 1961). As a highly territorial generalist predator, boobooks are probably more 2225

capable than most bird species of competing successfully for nest hollows, even under 2226

circumstances where suitable hollows are limited and competition from introduced species 2227

is high. If this hypothesis is correct, supplemental provision of artificial nest hollows would 2228

not be expected to increase boobook abundance unless suitable hollows are nearly absent 2229

from an area. 2230

Care should be taken not to generalise this conclusion to all predatory species 2231

utilizing tree hollows as nest sites. Nest competition has been suggested as a possible factor 2232

contributing to the decline of Norfolk Island boobooks and severe nest hollow limitation 2233

resulting from extensive habitat loss may have played a role in their decline (Olsen, 1996). 2234

Additionally, evolution in isolation from serious competition may have reduced the capacity 2235

of this subspecies to resist introduced aggressive mainland nest hollow competitors. Hollow 2236

availability may also vary with the size of the bird and the size of the hollow required for 2237

nesting. Powerful Owls are substantially larger than boobooks, take larger prey, and are 2238

potentially even less likely to be impacted by competition for nest hollows. However, their 2239

requirement for larger nest hollows – which are often scarcer in fragmented habitats – has 2240

apparently led to failures of established pairs to breed until a suitable nest box was provided 2241

(Isaac et al., 2014a). 2242

While boobooks and other predatory birds are unlikely to be severely impacted by 2243

nest competition by most introduced bird species, they may be negatively impacted by 2244

other potential nest hollow competitors. Colonization of nest hollows of all sizes by feral 2245

honeybees has been noted to be particularly problematic in southwest Western Australia 2246

(Johnstone et al. 2015) and colonization of boobook hollows by feral bees (Johnstone and 2247

Kirkby, 2007) was specifically recorded. In some instances, owls that have apparently been 2248

stung to death by bees have been observed in hollows (Western Australian Museum, n.d.), 2249

suggesting that feral honey bee nest competition sometimes directly contributes to 2250

boobook mortality. It is suspected that, in Western Australia, the impact of feral bees on 2251

cavity nesting birds is greatest in the Wheatbelt where canola crops prompt more frequent 2252

swarming (Johnstone and Kirkby 2007). Nest boxes used in our study incorporated several 2253

features intended to deter use by feral honeybees and our results may not be 2254

119

representative of honeybee competition rates for all nest box designs or in natural hollows. 2255

Anecdotally, we observed honeybees using several nest boxes unrelated to our study in the 2256

Perth metropolitan area. We encourage authors of future studies involving nest boxes to 2257

carefully report on all aspects of nest box design as this is an important and frequently 2258

overlooked factor impacting life history parameters of animals using the boxes (Lambrechts 2259

et al., 2012). 2260

Conclusion 2261

While artificial nest boxes are important tools in wildlife research and conservation, 2262

our study indicates that their use is not a panacea for every situation where hollow nesting 2263

species are in need of conservation management. In an era of heavily constrained 2264

conservation budgets, ineffective nest boxes intended to improve abundance of 2265

conservation-dependant species may divert valuable funds from more effective uses. 2266

Furthermore, poor design or application in inappropriate circumstances may lead to 2267

unintended negative outcomes for native biodiversity or unanticipated bias in scientific 2268

studies. 2269

Acknowledgments 2270

This project was supported financially by The Holsworth Wildlife Research 2271

Endowment via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird 2272

Research Award, and the Edith Cowan University School of Science Postgraduate Student 2273

Support Award. We thank the Western Australia Department of Biodiversity, Conservation, 2274

and Attractions, and the many city councils and private landowners who provided access to 2275

the sites involved in this project. This research was made possible by the generous 2276

assistance of dozens of volunteers who assisted in boobook surveys and nest monitoring. 2277

We especially thank Dr. Cheryl Lohr for providing valuable assistance in statistical analysis. 2278

2279

120

Chapter 6 Toxoplasma gondii seropositivity across urban and 2280

agricultural landscapes in an Australian owl 2281

2282

Lohr, M. T., C. A. Lohr, A. H. Burbidge, and R. A. Davis. Toxoplasma gondii seropositivity 2283

across urban and agicultural landscapes in an Australian owl. Veterinary Parasitology. 2284

(In Preparation). 2285

2286

Abstract 2287

Toxoplasma gondii is an apicomplexan parasite with a wide host range and 2288

cosmopolitan distribution. House cats (Felis catus) and other members of the family Felidae 2289

are the definitive hosts for T. gondii. Members of the family Felidae were absent from 2290

Australia until house cats were brought to the continent by European explorers and 2291

colonists and the lack of evolutionary history with T. gondii has been hypothesized to leave 2292

native Australian fauna more susceptible to the negative impacts of infection. As a 2293

consequence, understanding the factors that drive differences in environmental prevalence 2294

of T. gondii may inform conservation strategies for vulnerable Australian wildlife. As cat 2295

abundance has been documented to vary with landscape composition, we hypothesized 2296

that T. gondii infection would be more prevalent in urban and agricultural landscapes than 2297

landscapes dominated by intact bushland. The Australian Boobook (Ninox boobook) was 2298

used as a test species because it has been suggested that non-migratory owls may be useful 2299

indicators of ecosystem wide T. gondii contamination. We used modified agglutination tests 2300

to determine seropositivity in serum and meat juice samples from boobooks across 2301

landscapes dominated by urban/periurban development, agriculture and intact bushland. 2302

We also examined correlations between T. gondii seropositivity and other factors like age, 2303

season, injury status, and exposure to environmental pollutants which could impact 2304

likelihood of infection. Moderately low levels of seropositivity were detected across all 2305

samples. We believe that this is the first published instance of T. gondii seropositivity in a 2306

wild predatory bird in Australia. Most risk factors previously implicated in increased risk of 2307

T. gondii infection did not show significant correlations with observed seropositivity in 2308

boobooks. However, the season in which the sample was collected did correlate 2309

121

significantly with seropositivity. We suggest that seasonally dependant environmental 2310

factors which influence oocyst viability may obscure any relationship between landscape 2311

type and latent T. gondii infection rates in boobooks. 2312

Introduction 2313

Parasites and pathogens are increasingly implicated as a contributing factor in the 2314

declines of wildlife across the globe but a lack of baseline data has complicated efforts to 2315

understand the impacts of specific organisms (Pacioni et al. 2015). Well-documented severe 2316

impacts of parasitic organisms include worldwide declines in amphibian populations due to 2317

chytrid fungus (Batrachochytrium dendrobatidis) infection (Houlahan et al. 2000), ongoing 2318

reductions in North American bat populations as a result of white nose syndrome caused by 2319

the fungus Geomyces destructans (Foley et al. 2011), and the extinction or decline of most 2320

Hawaiian honeycreeper species due to avian malaria (Plasmodium relictum) (Warner, 1968). 2321

While introducing novel parasites to immunologically naïve hosts can have potentially 2322

devastating consequences, the impacts of parasites on their hosts are often more subtle 2323

(Pacioni et al. 2015). Similarly, while managing parasites and pathogens may be the key to 2324

some conservation efforts, there are very few studies on avian parasite ecology in Australia 2325

(Delgado-V. and French, 2012; Ford et al., 2001) with authors noting that “Virtually nothing 2326

is known about the effect of disease and parasites on Australian birds” (Ford et al. 2001). 2327

If parasitism is a threatening process for Australian birds, it may be exacerbated by 2328

anthropogenic land uses, sampling methodologies, or other threatening processes such as 2329

anticoagulant rodenticides (ARs) (Lemus et al., 2011; Riley et al., 2007; Serieys et al., 2018). 2330

To fully understand the implications of parasites and pathogens on avian conservation 2331

efforts it is necessary to examine patterns of parasitism across multiple habitat types. For 2332

example, Cooper’s Hawks (Accipiter cooperi) appeared to preferentially inhabit urban areas 2333

of Tucson Arizona (Battin, 2004). Urban environments tend to maintain high densities of 2334

prey species which may serve as a cue for habitat selection (Isaac et al. 2014). However, 2335

Boal & Mannan (1999) observed higher rates of nest failure in Cooper’s Hawks in urban 2336

areas than in periurban areas, due to nestlings being killed by trichomoniasis. This disease is 2337

caused by a protozoan vectored by feral pigeons which are abundant in urban areas and 2338

made up a much higher proportion of the Cooper’s Hawks’ diets in urban areas (Boal and 2339

Mannan, 1999). Hence, Cooper’s Hawks were being drawn out of less anthropogenically-2340

122

altered habitats into areas with higher prey abundance and higher risk of parasitic infection 2341

which reduced their fecundity to unsustainable levels. Documentation of demographic 2342

parameters within the city that did not explain the population’s stability indicates that the 2343

population was sustained by migration from outside urban areas and that parasites in the 2344

urban area had created an ecological trap (Battin, 2004). Similarly, birds in urban areas in 2345

Brazil were found to have higher infection rates of haemosporidian blood parasites than 2346

birds in intact natural landscapes (Belo et al., 2011). Reduced parasitism in urban areas has 2347

also been observed in some bird species but the direction of the trend is probably 2348

dependant on the type of parasite and its mode of transmission (Delgado-V. and French, 2349

2012; Suri et al., 2016). Evaluating the impact of landscape-level human land use practices 2350

on parasite prevalence will be increasingly important to the conservation of native fauna as 2351

the area of land subject to urban and agricultural development increases. 2352

Effects of Toxoplasma gondii on Humans and Wildlife 2353

Toxoplasma gondii is a parasitic protozoan capable of infecting a wide taxonomic 2354

range of warm-blooded vertebrates. Felids are its only known definitive hosts (Miller et al. 2355

1972). T. gondii causes both acute and latent toxoplasmosis (Remington and Cavanaugh, 2356

1965). It is known for its capacity for manipulation of host behaviour and increasing 2357

susceptibility to predation by cats by increasing dopamine metabolism in the brain of 2358

infected secondary hosts (Prandovszky et al. 2011). In humans, acute toxoplasmosis can 2359

cause severe illness in newborns and immunocompromised individuals and can cause 2360

spontaneous abortion and foetal abnormalities (Wolf et al. 1939). However, consensus is 2361

emerging among medical professionals that latent toxoplasmosis is not benign. It has been 2362

implicated as a risk factor in a number of serious health problems including epilepsy 2363

(Ngoungou et al. 2015), generalized anxiety disorder (Markovitz et al. 2015), schizophrenia 2364

(Torrey et al. 2007), impaired reaction time (Havlícek et al. 2001), car accidents (Flegr et al. 2365

2002), obsessive compulsive disorder (Miman et al. 2010b), Parkinson’s disease (Miman et 2366

al. 2010a), Alzheimer’s disease (Kusbeci et al. 2011), Down Syndrome (Prandota, 2010), and 2367

suicide attempts (Arling et al. 2009). Worldwide correlations with other diseases were 2368

examined by Flegr et al. (2014). It has even been proposed that T. gondii may have a 2369

worldwide impact on human culture by subtly altering the neurochemistry of substantial 2370

proportions of the global population (Lafferty, 2006). Because roughly one third of the 2371

123

worldwide human population is believed to be infected with T. gondii, there is growing 2372

concern over the impacts of this organism on global human health. 2373

T. gondii infection is also a concern for native Australian wildlife. Lack of a felid 2374

definitive host prior to the arrival of Europeans suggests that native wildlife do not share a 2375

long evolutionary history with T. gondii. Acute toxoplasmosis has been observed to be 2376

lethal in a wide variety of Australian native marsupial species and can cause blindness, 2377

lethargy, respiratory and digestive problems, and decreased coordination (Patton et al. 2378

1986; Canfield et al. 1990). T. gondii infection likely increases the probability of predation in 2379

some native mammals (Obendorf et al. 1996) and has been implicated in the decline of 2380

some wild marsupial populations but the extent of its impact is not well understood 2381

(Freeland, 1994). Three macropod species were observed to be infected with multiple 2382

strains of T. gondii (Pan et al. 2012) and cat predation coupled with T. gondii infection may 2383

have played a role in the observed local decline of another macropod species (Fancourt, 2384

2014). Death by acute toxoplasmosis has also been observed in ten species of native 2385

Australian birds held in captivity (Hartley and Dubey, 1991) including one penguin which had 2386

only been in captivity for a few days (Mason et al., 1991) but prevalence of T. gondii 2387

infection has not been quantified in wild bird populations and has not been examined at a 2388

landscape level. 2389

Predatory Birds and Toxoplasma gondii Infection 2390

Toxoplasma infection has been observed in owls and other raptors in the wild in 2391

North America and Europe (Kirkpatrick et al. 1990; Lindsay et al. 1993; Dubey et al. 2010; 2392

Lopes et al. 2011; Yu et al. 2013), with some species documented to have seroprevalence 2393

rates of nearly 80% (Aubert et al. 2008). Prevalence of T. gondii infection in wildlife is a good 2394

indicator of environmental contamination by oocysts and is useful in assessing risk to 2395

human health (Dubey and Jones, 2008). Similarly, T. gondii infection is more likely to be 2396

detected in predators which typically have higher rates of seroprevalence than omnivores 2397

and herbivores as a result of greater risk of ingesting infected animals over their lifetime 2398

(Hejlícek et al., 1997; Hollings et al., 2013). Birds in particular are preferable as 2399

environmental bio-monitors for T. gondii because vertical transmission (direct congenital 2400

transmission from adult to offspring) has not been observed in Australian birds, in contrast 2401

to marsupials (Parameswaran et al. 2009). Vertical transmission in a population could make 2402

124

seroprevalence rates a less useful index of overall environmental contamination. Vertical 2403

transmission is unlikely in birds due to the extremely low incidence of T. gondii in eggs 2404

(Dubey, 2010). Many raptor species prey primarily on small birds and mammals that are 2405

frequently intermediate hosts of T. gondii and, as such, raptor seroprevalence rates have 2406

the potential to offer valuable insights into environmental prevalence of T. gondii (Love et 2407

al. 2016). 2408

At present, seroprevalence rates in wild Australian raptors are unknown and within-2409

species differences across different land-use categories have not been studied in any raptor 2410

species worldwide. Non-migratory owl species, such as Australian Boobooks (Ninox 2411

boobook), have been specifically identified as useful indicator species (Gazzonis et al., 2018) 2412

for assessing differences in environmental T. gondii oocyst contamination on a landscape 2413

scale. T. gondii is known to infect owls (Dubey et al., 1992) but direct mortality and 2414

observable illness resulting from infection are extremely uncommon (Mikaelian et al., 1997). 2415

Sub-lethal effects of T. gondii infection on owls are largely unknown but most carnivorous 2416

birds are assumed not to be affected by acute clinical toxoplasmosis (Dubey et al., 2010; 2417

Love et al., 2016). The ability to become infected without obvious signs of increased 2418

mortality rates is a desirable attribute in effective bio-monitors. Boobooks are an ideal 2419

species for monitoring landscape-level T. gondii prevalence because they are found in a 2420

wide range of habitat types across Australia – including those that have been substantially 2421

altered by humans. 2422

Aims 2423

We sought to determine whether correlations exist between different types of 2424

human land use and T. gondii infection rates in raptors. Domestic cat density has been 2425

observed to correlate strongly with housing density (Sims et al. 2008) and spatial correlation 2426

between T. gondii seropositivity and both human habitation and cat density has been noted 2427

in carnivorous wildlife (Barros et al., 2018; Hollings et al., 2013). To identify the relative 2428

importance of landscape type in risk of T. gondii infection, we also assessed other factors 2429

associated with increases in toxoplasma seropositivity in wildlife including age (Cabezón et 2430

al., 2011; Lindsay et al., 1993; Lopes et al., 2011), injury status (Hollings et al., 2013), and 2431

sampling during seasons with more favourable conditions for T. gondii oocyst survival 2432

(Simon et al., 2018). We also explored potential associations between T. gondii 2433

125

seropositivity and anticoagulant rodenticides (ARs) because an emerging body of research 2434

has shown correlations between ARs and parasites and infectious diseases (Lemus et al., 2435

2011; Riley et al., 2007; Serieys et al., 2015; Vidal et al., 2009). 2436

Consequently, we expected to observe the highest rate of boobook seropositivity in 2437

urban and periurban areas, where both cats and commensal rodents exist in elevated 2438

densities as a result of human activities. We hypothesized that: 2439

1) Seroprevalence in agricultural landscapes would be intermediate between 2440

seroprevalence observed in primarily urban/periurban landscapes and landscapes 2441

dominated by native bushland. 2442

2) T. gondii seropositivity will be higher in individuals which are older, sampled 2443

during wetter seasons, and in the dead or injured category due to increased reaction time 2444

associated with infection potentially increasing the risk of collisions with windows and 2445

motor vehicles. 2446

3) T. gondii seropositivity will be higher in birds that have detectable levels of 2447

anticoagulant rodenticides (ARs). 2448

Methods 2449

Sample Collection 2450

We used several methodologies to collect boobook blood and tissue samples across 2451

a variety of habitat types present in Western Australia, with an active focus on procuring 2452

samples in the Perth Metropolitan Area, areas of intensive agriculture within an 2453

approximate 60km radius of the town of Kellerberrin in the central wheatbelt approximately 2454

200 km east of Perth, and intact forested areas of the Perth Hills between the two types of 2455

fragmented landscape. During occupancy surveys for another study (Chapter 5), boobooks 2456

were located at night across all three habitat types using recorded boobook calls broadcast 2457

on a portable speaker. Additionally, boobooks roosting during the day were located with 2458

the assistance of volunteers and were also opportunistically included in the study. Wild 2459

boobooks were captured using a noose pole similar to methodology used to capture 2460

boobooks elsewhere (Olsen et al., 2011). After banding and basic biometric measurements, 2461

boobooks were assigned to age classes of one year or less ('hatch year') or greater than one 2462

126

year ('after hatch year') based on the presence of juvenile down and by examination of 2463

fluorescence patterns in the undersides of the remiges under ultraviolet light (Weidensaul 2464

et al., 2011). In some instances, age class could not be assigned because of degradation of 2465

porphyrins by exposure to sunlight. Following these measurements, 0.5 ml of blood was 2466

taken from the jugular vein of all boobooks. Live boobooks held by wildlife rehabilitators 2467

were also sampled if the bird was sufficiently healthy for blood sampling. Blood was taken 2468

from the right jugular vein to reduce handling time and risk of hematoma relative to 2469

sampling from the brachial vein (Owen, 2011). Blood taken from these boobooks was 2470

allowed to coagulate for approximately 24 hours. Samples were then centrifuged for 10 2471

minutes at 13,200 RPM to produce serum. Serum was stored at -20°C until testing. 2472

Boobooks found dead by volunteers or euthanized by wildlife rehabilitators were 2473

opportunistically sampled as well (see methods in Lohr, 2018). Heart and breast muscle 2474

tissue were removed from dead boobooks which were not in an advanced state of 2475

decomposition. These tissues were placed in a plastic bag, and stored frozen at -20°C. Prior 2476

to testing, the specimens were thawed and 0.5ml of resulting fluids (hereafter “meat juice”) 2477

was removed from the bag using a syringe. Meat juice samples were then centrifuged for 2478

10 minutes at 13,200 RPM and the supernatant was removed and stored at 4°C until it was 2479

used for testing within 24 hours of tissue thawing. 2480

Serological Testing 2481

Serum and meat juice were both tested using a commercially available modified 2482

agglutination test (Toxo-Screen DA, BioMerieux, France). Modified agglutination tests 2483

(MATs) are the preferred serologic tests used in detecting chronic toxoplasma infection in 2484

wild birds because they are sensitive, specific, do not require special equipment, and appear 2485

to work well across all avian species tested (Dubey, 2002). Testing was conducted according 2486

to the instructions included with the commercial kit. The only variation from the testing kit 2487

instructions was that we used serum dilutions of 1:25 and 1:400 and meat juice dilutions of 2488

1:4 and 1:64. This synchronized maximum concentrations with those used in previous 2489

testing of similar matrices (Cabezón et al., 2011; El-Massry et al., 2000) and (Richomme et 2490

al., 2010), respectively) using the same commercially available testing kit and maintained 2491

equal dilution ratios between the two matrix types we tested. A higher concentration of 2492

meat juice was used in testing because meat juice contains lower concentrations of 2493

127

Toxoplasma antibodies than serum (Richomme et al., 2010). In accordance with previous 2494

literature, we considered positive results in samples diluted at a ratio of 1:25 as indicative of 2495

latent Toxoplasma infection in serum samples (Aubert et al., 2008; El-Massry et al., 2000) 2496

and used a threshold of 1:4 for determining seropositivity in meat juice samples (Richomme 2497

et al., 2010). Positive and negative controls provided with the kit were included in each 2498

testing plate. A total of 130 individuals were tested. Of these, 61 were tested using only 2499

serum, 61 were tested using only meat juice and eight were tested using both serum and 2500

meat juice. 2501

In eight instances, both serum and meat juice were available for the same individual 2502

birds due to either subsequent euthanasia of boobooks held in care which failed to recover 2503

sufficiently to allow release or banded boobooks being handed in by members of the public 2504

after being found dead. 2505

2506

Statistical Analysis 2507

We used RStudio 1.1.383 (RStudio, Inc., Boston,MA, USA) to conduct Fisher's exact 2508

tests in order to examine correlations between T. gondii seropositivity and a number of 2509

potentially relevant environmental and demographic variables because in all tests the 2510

number of observations in at least one category was ≤ 5 (Gazzonis et al., 2018). Tested 2511

variables included landscape type (agriculture, bushland, urban/periurban), the age of the 2512

boobook sampled (hatch year or after hatch year), season in which the sample was collected 2513

(winter, spring, summer, autumn), the status of the boobook when sampled (wild or 2514

compromised (in care or dead)). Sample sizes varied slightly between the individual 2515

statistical tests because, in some cases, volunteers provided incomplete collection 2516

information regarding collection date and location or because we were unable to accurately 2517

determine the age of the boobook. This necessitated the exclusion of some deceased 2518

individuals from particular statistical tests. 2519

A subset of deceased boobooks (N=65) tested for AR residues in a previous study 2520

(Lohr, 2018) were used to test for correlations between T. gondii seropositivity and total 2521

concentrations of ARs in liver tissue. Individuals were assigned to four categories of AR 2522

exposure based on biologically relevant thresholds. The lowest category included all 2523

128

samples with total AR concentration below 0.01 mg/kg because it is the limit of 2524

quantification for most of the ARs tested (Lohr, 2018). Additional thresholds of 0.10 mg/kg 2525

and 0.50 mg/kg were also used to delineate the remaining three categories because 0.10 2526

mg/kg is regularly used as the lower limit for potential toxic effects in raptors (Albert et al., 2527

2010; Christensen et al., 2012; Langford et al., 2013; Ruiz-Suárez et al., 2014; Shore et al., 2528

2016; Stansley et al., 2014; Walker et al., 2011, 2008) and liver concentrations of 0.50 mg/kg 2529

are likely to be lethal in most birds (Dowding et al., 1999). We used Fisher’s exact test to 2530

determine if toxoplasma seropositivity was associated with AR exposure. 2531

Results 2532

In the eight boobooks with both serum and blood samples available, six were 2533

negative in both samples and two appeared to seroconvert and had negative serum samples 2534

but positive meat juice samples. Across all 130 boobooks sampled, 13.1% were seropositive 2535

for T. gondii in at least one sample. Seropositivity was more prevalent in meat juice samples 2536

(18.0% n = 61) than in serum samples (6.6% n = 61) but did not differ significantly between 2537

the two matrix types sampled (P=0.096). Consequently, for analyses other than direct 2538

comparisons between the serum and meat juice seropositivity and for comparisons 2539

involving AR exposure which was only testable in dead boobooks, the data from both 2540

matrices were pooled and boobooks testing positive in either matrix were treated as 2541

positive samples. The only factor which significantly correlated with T. gondii seropositivity 2542

was the season in which the sample was collected (p = 0.024) (Error! Reference source not 2543

ound.). While subsequent pairwise comparisons between seropositivity by season were not 2544

significant, seropositivity rates were numerically higher in autumn and winter relative to 2545

spring and summer (Error! Reference source not found.) and the difference between 2546

ummer and autumn seropositivity rates was marginally non-significant (p = 0.099). Overall 2547

anticoagulant rodenticide exposure did not show significant associations with T. gondii 2548

seropositivity but seroprevalence was numerically lower in boobooks with total liver 2549

concentrations of ≤ 0.01mg/kg (13.0%) than in the three higher categories (23.7% to 25%) 2550

(Figure 6.2). 2551

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

Variable Category Positive/ examined Seroprevalence (%) p-value

Testing matrix serum 4/61 6.6 0.096

129

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

130

2555

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

2558

131

2559

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

Discussion 2563

Apparent seroconversion in the two individuals which tested negative in serum 2564

samples but positive in meat juice samples may be an artefact of the MAT test used. False 2565

negative results can be obtained during acute stages of T. gondii infection because the test 2566

is only sensitive to IgG antibodies, and not IgM antibodies which are present at the onset of 2567

infection (Sroka et al., 2008). It is entirely plausible that the two boobooks were 2568

experiencing active infections when initially sampled but their infections would only be 2569

detected by the subsequent meat juice sampling. We believe that this explanation, in 2570

combination with the lack of a significant difference in seropositivity rates between serum 2571

and meat juice samples justifies our decision to combine data from both matrix types in the 2572

other analyses. 2573

132

Studies using the same commercially available modified agglutination test on other 2574

continents have found varying rates of seropositivity across multiple raptor species: 34.5% 2575

in the south-eastern USA (n=281) (Love et al., 2016), 25.7% in Taiwan (n=206) (Chen et al., 2576

2015), 35.8% in France (n=53) (Aubert et al., 2008), 50.0% in Portugal (Lopes et al., 2011), 2577

and 10.7% in Italy (n=93) (Gazzonis et al., 2018). While at the lower end of the scale, our 2578

results (13.1%) were within the ranges previously reported in studies of predatory birds. 2579

Interestingly, overall seropositivity was nearly identical to the rate of 13.0% reported for a 2580

native marsupial carnivore (chuditch, Dasyurus geoffroii), in Julimar Valley (Parameswaran, 2581

2008), an area of continuous bushland adjacent to our sites in the Perth Hills. 2582

Four potential explanations exist for the relatively low seropositivity rates we 2583

observed. Australian boobooks have not previously been evaluated using this test and it is 2584

possible that species-specific factors may have led to false negative results. We view this 2585

scenario as unlikely because, while false negatives using MAT are common in some species – 2586

particularly in dogs (Liu et al., 2015) – this test has been used successfully to detect 2587

toxoplasma seropositivity in a wide variety of other predatory bird species (Chen et al., 2588

2015; Gazzonis et al., 2018; Lopes et al., 2011). 2589

It is also unlikely that our use of meat juice in addition to serum would have reduced 2590

detections relative to other studies. A study directly examining detection of T. gondii 2591

antibodies in meat juice did not find reduced detectability or degradation of antibodies in 2592

response to repeated freezing and thawing of meat (Mecca et al., 2011). If anything, the 2593

use of this methodology should have increased seropositivity detection in our study relative 2594

to other studies which tested only serum. The numerically but not significantly higher 2595

detection rate of T. gondii antibodies in meat juice samples is consistent with this 2596

hypothesis. 2597

The diet and trophic position of boobooks may also provide some explanation for 2598

the relatively low seropositivity rates seen in boobooks. Australian Boobooks are medium-2599

sized owls (Olsen, 2011a) and consume a variety of invertebrate and vertebrate prey (Trost 2600

et al., 2008). A study on T. gondii seropositivity in wild birds in Spain found seropositivity 2601

rates ranging from 0% to 25% in six small and medium sized owl species (Cabezón et al., 2602

2011). However, the same study detected T. gondii antibodies in 68% of all Eurasian Eagle-2603

133

owls (Bubo bubo) (Cabezón et al., 2011). This species is substantially larger and occupies a 2604

higher trophic level than the other owl species tested. In the context of this research, the 2605

seropositivity of boobooks we observed is typical of an owl species of its size and diet. 2606

Landscape Type 2607

The relatively warm and dry climate in our study areas may explain our observation 2608

of lower seropositivity rates than in most other areas of the world where raptors have been 2609

sampled. Worldwide, toxoplasma prevalence is lowest in hot arid areas, presumably due to 2610

shorter duration of oocyst viability under hot dry conditions (Meerburg and Kijlstra, 2009). 2611

In a study examining habitat impacts on T. gondii seropositivity in wild rabbits, seropositivity 2612

was substantially higher in habitats with more shade and humidity (Almería et al., 2004). 2613

This pattern may explain the lack of significant difference observed in seropositivity 2614

between landscape types. Counter-intuitively, clearing of land for urban and agricultural 2615

uses could lead to a reduction in T. gondii seroprevalence despite potential increases in cat 2616

abundance if the reduction in vegetative cover results in an increase in soil temperature and 2617

decrease in soil moisture, leading to inhibition of T. gondii oocyst viability. Future work 2618

examining T. gondii seroprevalence in a single intermediate host species across paired 2619

habitat types in areas with substantially different rainfall levels would be useful in 2620

determining relative contributions of cat abundance and soil moisture to seropositivity in 2621

intermediate hosts. 2622

Age 2623

We were surprised that no difference in seropositivity was detected between age 2624

classes. Some studies have found that T. gondii detection increases with age in wild 2625

predatory birds (Cabezón et al., 2011; Lindsay et al., 1993; Lopes et al., 2011) which is in 2626

keeping with the hypothesized lifelong persistence of the parasite after infection. However, 2627

other studies of predatory birds (Gazzonis et al., 2018) and wild rabbits (Oryctolagus 2628

cuniculus) (Almería et al., 2004) have failed to detect a difference in seropositivity between 2629

different age classes but did not address why no correlation was detected. It is possible that 2630

our grouping of boobooks into two coarse age classes of < one year and ≥ one year obscured 2631

longer-term trends in seropositivity. Because boobooks are relatively long lived – one was 2632

re-sighted in the field alive nearly 16 years after it was originally banded (Commonwealth of 2633

134

Australia, 2015) – they are potentially at risk from the long-term effects of latent T. gondii 2634

infection similar to those observed in humans. Impaired reaction time resulting from T. 2635

gondii infection is cumulative in humans and increases with the duration of latent infection 2636

(Havlícek et al. 2001). Similar effects in long-lived wildlife could predispose individuals to 2637

greater risk of predation, accident, or vehicular collision. Use of predatory bird species with 2638

a greater number of more easily-identifiable age categories (such as Wedge-tailed Eagles 2639

(Aquila audax)) could help to resolve questions relating to both changes in seropositivity 2640

between age classes and whether cumulative impacts of latent toxoplasmosis are 2641

problematic for predatory birds. 2642

Injury Status 2643

The lack of significant difference in seropositivity between boobooks found dead or 2644

held by wildlife carers and those captured in the wild was also unexpected and runs 2645

contrary to observations by Hollings et al. (2013) in Tasmanian pademelons (Thylogale 2646

billardierii) shot for pest control purposes and pademelons killed by motor vehicle collisions. 2647

It is unlikely that this is a consequence of our testing of multiple matrix types, as 2648

seropositivity was numerically – though not significantly – higher in meat juice samples from 2649

deceased boobooks. It seems more likely that our inclusion of boobooks which were killed 2650

or disabled by a wide variety of causes may have obscured any potential effect specific to 2651

motor vehicle collisions. In humans, reduced concentration time and increased reaction 2652

time were proposed as the mechanisms by which T. gondii seropositivity increased rates of 2653

car accident (Flegr et al., 2002). It is unlikely that these potential causative factors are 2654

relevant to all the causes of mortality or injury associated with the boobooks in our study. 2655

Unfortunately, uncertainty over proximate causes of death or injury in the boobooks we 2656

tested precluded direct testing of a more specific relationship with seropositivity. 2657

Season 2658

Several potentially interacting biological factors could plausibly explain the higher 2659

seropositivity rates of boobook samples obtained in autumn and winter. Boobooks 2660

consume a higher proportion of mammals and birds in winter relative to other times of year 2661

when insects make up a larger percentage of their diet (Trost et al., 2008). Additionally, 2662

temperature and rainfall patterns in autumn and winter in southwest Western Australia are 2663

more conducive to T. gondii oocyst viability and, as a consequence, infection rates in prey 2664

135

species may be higher at this time of year. A similar explanation was given for observations 2665

of higher T. gondii seroconversion rates observed in house cats in autumn and winter 2666

relative to spring and summer (Simon et al., 2018). Both factors may increase the risk of 2667

boobooks consuming prey with tissue cysts containing T. gondii bradyzoites and subsequent 2668

infection in winter. Boobooks with recent infections may also have been easier to capture 2669

or more likely to die and be injured and consequently be sampled by our study, increasing 2670

the detected seropositivity of individuals sampled at this time of year. 2671

Anticoagulant Rodenticide Exposure 2672

Alternately, exposure to anticoagulant rodenticides may have contributed to 2673

increased seropositivity of samples obtained in winter. Significantly higher liver 2674

concentrations of anticoagulant rodenticides have been observed in boobooks in the Perth 2675

metropolitan area in winter relative to spring and summer (Lohr, 2018). Additionally, while 2676

the Fisher’s exact test failed to detect a significant difference between rodenticide exposure 2677

categories, the seroprevalence of boobooks in the lowest exposure category with 2678

insubstantial amounts of rodenticide was numerically lower than the three categories with 2679

clinically relevant AR exposure (≥0.01 mg/kg) (Figure 6.2). Sub-lethal exposure to 2680

anticoagulant rodenticides has been found to correlate with immune dysfunction in bobcats 2681

(Lynx rufus) (Serieys et al., 2018) and has been hypothesized as the explanation for an 2682

observed correlation between anticoagulant rodenticides and notoedric mange (Riley et al., 2683

2007). These correlations are to some degree called into question by a study on domestic 2684

cats (Felis catus) which did not find a substantial link between anticoagulant rodenticides 2685

and immune dysfunction (Kopanke et al., 2018). However, even if immunosuppression is 2686

not the mechanism by which AR exposure facilitates hyper-parasitism, similar increases in 2687

pathogen and parasite load correlated with AR exposure have also been documented in 2688

Great Bustards (Otis tarda) exposed to the AR chlorophacinone (Lemus et al., 2011). 2689

If AR exposure facilitated reactivation of latent toxoplasmosis, this could explain the 2690

increase in seroprevalence detected in winter and autumn. Alternately, it is possible that a 2691

synergistic interaction between AR exposure and T. gondii infection increased the 2692

probability of the boobooks dying and entering this study to be tested. A synergistic effect 2693

on probability of mortality involving the AR chlorophacinone and the pathogen Francisella 2694

tularensis has been suggested in common voles (Microtus arvalis) (Vidal et al., 2009). 2695

136

Another hypothesis potentially explaining the possible correlation between AR exposure 2696

and seropositivity is that boobooks could simply be exposed to both T. gondii and ARs at 2697

higher rates in winter months as a consequence of rodents making up a higher proportion of 2698

their diet at this time of year. These hypotheses are not mutually exclusive and would be 2699

difficult to distinguish without experimental study of this dynamic in a laboratory setting. 2700

While boobooks showed few significant trends in T. gondii seropositivity, this may be 2701

primarily an issue of low statistical power to detect such trends caused by relatively low 2702

sample sizes. This is a common problem when studying cryptic, nocturnal, carnivores which 2703

occurr at low densities and are difficult to capture. However, seasonal differences in 2704

seropositivity suggest that conditions influencing oocyst viability may be a more important 2705

determinant of exposure risk than the factors we examined directly. Future work evaluating 2706

the utility of boobooks and other raptors as bioindicators of environmental T. gondii 2707

contamination should examine seropositivity rates across temperature and rainfall 2708

gradients. The use of boobooks as bioindicators could help identify important landscape-2709

level drivers of T. gondii prevalence and has the potential to inform management actions 2710

and translocation efforts intended to benefit susceptible native mammals. 2711

Acknowledgments 2712

This project was supported financially by The Holsworth Wildlife Research Endowment 2713

via The Ecological Society of Australia, the BirdLife Australia Stuart Leslie Bird Research 2714

Award, and the Edith Cowan University School of Science Postgraduate Student Support 2715

Award. We thank Annette Koenders, Adriana Botero, and Louise Pallant for advice and 2716

technical assistance in serology testing. Our research would not have been possible without 2717

contributions of samples and access to live birds provided by Kanyana Wildlife 2718

Rehabilitation, Native Animal Rescue, Native ARC, Nature Conservation Margaret River 2719

Region, Eagles Heritage Wildlife Centre, and many individual volunteers especially Simon 2720

Cherriman, Angela Febey, Amanda Payne, Stuart Payne, and Warren Goodwin. 2721

2722

137

Chapter 7 Summary, Synthesis, and Management Implications 2723

I examined four distinct potential threatening processes which I predicted had the 2724

potential to vary in magnitude of impact between habitats fragmented by urban and 2725

agricultural land uses. Australian Boobooks (Ninox boobok) did not appear to be 2726

substantially negatively impacted by lack of nest hollow availability, infection with 2727

Toxoplasma gondii, or genetic isolation in either landscape type. However, I did detect 2728

considerable exposure to anticoagulant rodenticides (ARs) associated with the use of 2729

habitats containing commercial and residential development. 2730

In this chapter, I highlight the most important findings of each chapter and 2731

contextualise their relevance to their respective fields outside of the specific system I 2732

studied. I also discuss the contribution of my research to the theoretical framework in 2733

which the impacts of habitat fragmentation are typically evaluated. I then suggest specific 2734

practical implications of my findings for management actions intended to maintain or 2735

increase native biodiversity in landscapes dominated by intensive human land uses. 2736

Summary of major findings: 2737

Objective 1. Critically review literature on anticoagulant rodenticide exposure in native 2738

wildlife in Australia to clarify its role as a threatening process. 2739

My review of the literature relating to anticoagulant rodenticides in Australia 2740

revealed widespread anecdotal accounts of both primary and secondary anticoagulant 2741

rodenticide (AR) poisoning among a taxonomically diverse group of non-target wildlife. Key 2742

differences between Australia and other developed nations were noted in the regulation of 2743

ARs. Most notably, second generation anticoagulant rodenticides (SGARs) are readily 2744

available for purchase without a license in Australia, unlike in the United States and Canada. 2745

Australia is also one of only two countries to allow the use of the first generation 2746

anticoagulant rodenticide (FGAR), pindone and to allow its use in widespread repeated 2747

baiting of natural systems for control, rather than eradication, of introduced species. 2748

Additional research is recommended to evaluate this practice. The review also identified 2749

patterns in world literature relating to reptiles and rodenticides which suggest the potential 2750

for high tolerance to rodenticides in at least some reptile taxa. Further experimental testing 2751

is necessary to determine if this hypothesized resistance makes reptiles efficacious vectors 2752

138

of ARs to humans and predatory wildlife. If so, rodenticide poisoning in warmer areas of the 2753

world with diverse and abundant reptile herpetofaunas, may be a greater threat to 2754

predatory wildlife than in the cool temperate regions where most AR ecotoxicology work 2755

has been conducted. 2756

Objective 2. Investigate the relationship between exposure to anticoagulant rodenticides 2757

and urban and agricultural fragmentation. 2758

Exposure to anticoagulant rodenticides (ARs) was prevalent in the boobooks tested 2759

(72.6%) and higher than typically observed in similar studies of predatory birds on other 2760

continents. The vast majority of the rodenticides detected were the more persistent second 2761

generation anticoagulant rodenticides (SGARs). AR exposure correlated positively with 2762

proximity to urban/periurban habitat at all spatial scales and negatively with use of 2763

agricultural areas and native bushland. The association between AR exposure and the 2764

proximity of boobooks to urban and suburban development (but not agricultural land uses), 2765

supports modelling which suggests that matrix type can exert strong influences on wildlife 2766

inside habitat patches (Sisk et al., 1997). The strongest correlations between AR exposure 2767

and habitat were found at the spatial scale of a boobook’s estimated home range. This 2768

suggests that predatory birds with larger home ranges may be at risk of AR exposure over a 2769

larger proportion of the landscape. Additional research on non-target AR exposure in 2770

Australia is urgently needed to determine the level of threat posed to other wildlife species, 2771

particularly carnivores and scavengers with large home ranges which are already listed as 2772

threatened (e. g. quolls (Dasyurus sp.) and Tasmanian devils (Sarcophilus harrisiii). 2773

Objective 3. Determine if urban and agricultural fragmentation influence boobook genetic 2774

structure. 2775

Boobooks did not exhibit substantial genetic structure among landscapes dominated 2776

by urban development, agricultural crops, or native bushland in between. This trend held 2777

with the inclusion of boobook samples originating across a larger geographic area including 2778

the majority of Western Australia. Banding data from my study and others demonstrated 2779

that fledgling boobooks are capable of dispersing across urban habitats for distances far 2780

greater than those between remaining bushland fragments. In combination, these findings 2781

suggest a high degree of landscape permeability and genetic connectivity in boobooks 2782

across all areas sampled. Highly mobile species have a greater probability of survival than 2783

139

less mobile species in areas which have experienced habitat fragmentation (Ewers and 2784

Didham, 2006). High mobility despite fragmentation coupled with the apparent capacity to 2785

use matrix habitat in at least some circumstances likely explains the persistence of 2786

boobooks in highly fragmented landscapes, albeit at lower densities. 2787

Objective 4. Examine whether nest box supplementation increases site occupancy at 2788

unoccupied sites and whether this effect differs between urban and agricultural landscapes. 2789

Boobooks occupied fewer sites in urban and agricultural remnant bushlands than in 2790

continuous woodland. Nest box supplementation at unoccupied sites did not alter site 2791

occupancy over the duration of this study. However, one nest box in an urban bushland 2792

remnant was successfully used by a boobook. Nest hollows do not appear to be a limiting 2793

factor in the use of remnant woodlands by boobooks in either fragmented landscape type 2794

despite boobooks being obligate hollow nesters. Nest box supplementation is unlikely to be 2795

an effective tool for increasing boobook abundance in remnant woodlands but anecdotal 2796

observations of boobooks utilising nest boxes in urban areas completely devoid of native 2797

bushland suggest that nest boxes may reduce matrix hostility and increase usable space in 2798

highly-altered areas lacking remaining suitable tree hollows. 2799

Objective 5. Explore patterns of Toxoplasma gondii seropositivity in boobooks across the 2800

urban, agricultural, and natural landscapes. 2801

Toxoplasma gondii seropositivity did not vary significantly among urban, agricultural, 2802

and woodland dominated landscape types. Most other factors which other studies have 2803

found to correlate with T. gondii seropositivity (i.e. age, season, injury status, and exposure 2804

to environmental pollutants) did not show significant correlations. Failure to detect these 2805

trends may have been caused by insufficient statistical power associated with low 2806

seropositivity rates. However, higher seropositivity was observed in cooler wetter seasons. 2807

This trend could be related to environmental conditions favouring oocyst viability, greater 2808

availability of infected prey, seasonal dietary shifts toward increase proportional 2809

consumption of prey species likely to be infected, or a combination of these factors. 2810

Increased risk of boobooks being infected by T. gondii associated with increased numbers of 2811

house cats (the definitive hosts for T. gondii) in urban and agricultural landscapes may be 2812

offset by decreased viability of oocysts in soil due to increases in soil temperature and 2813

decreases in soil moisture relative to areas of natural vegetation. This chapter reports what I 2814

140

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

in Australia. 2816

Synthesis 2817

My genetic research, in combination with band return data, revealed a biological 2818

reality more complex than the initial premise underpinning my research. I had initially set 2819

out to determine whether potential threatening processes operated differently in 2820

landscapes dominated by different anthropogenic matrices (i.e. urban and agricultural land 2821

use). Habitat fragmentation defined as “the process of subdividing a continuous habitat 2822

into smaller pieces” (Andrén, 1994) clearly occurred in landscapes consisting of 2823

predominantly urban or agricultural land use examined in this study, if pre-existing natural 2824

vegetation is considered to be synonymous with “habitat”. However, the lack of observed 2825

spatial genetic structure, dispersal of boobooks across urban matrix, active use of the 2826

agricultural matrix, and observed breeding inside highly developed urban areas with no 2827

adjacent bushland all suggest that boobooks are using human-dominated land cover types 2828

as well as remnant bushlands. If “habitat” is defined as “the subset of physical 2829

environmental factors that a species requires for its survival and reproduction” (Block and 2830

Brennan, 1993) then, by this definition there has been no “habitat fragmentation”. 2831

Essentially, from the perspective of boobooks, functional reduction in available habitat may 2832

not have occurred in landscapes dominated by agriculture and urban development despite 2833

extensive conversion of natural vegetation types. The misuse of the term “habitat” to mean 2834

something akin to “vegetation association” is common in published scientific literature 2835

(Franklin et al., 2002; Hall et al., 1997). The unresolved ambiguity and continuing misuse of 2836

the term “habitat” has led to the coining of the largely synonymous term “usable space” as 2837

an “area with habitat compatible with the physical, behavioral, and physiological 2838

adaptations of [an organism] in a time-unlimited sense” (Guthery et al., 2005). 2839

In the instance of boobooks examined in this study, true fragmentation of ‘usable 2840

space’” does not appear to have occurred across all areas of urban and agricultural land use. 2841

Boobooks were observed successfully fledging chicks in an area >3km from the nearest 2842

remaining patch of native vegetation and foraging in agricultural areas >1km from the 2843

nearest bushland, tree line, or patch of native vegetation (Chapter4). These behavioural 2844

observations are not conclusive but are strongly indicative that at least urban areas 2845

141

constitute ‘usable space’. Despite the apparent capacity of boobooks to successfully use 2846

urban and agricultural landscapes, the conversion of native vegetation associations to urban 2847

and agricultural land uses appears to have caused habitat degradation. This is supported by 2848

our observation of lower boobook occupancy rates in urban and agricultural bushland 2849

remnants, relative to areas of intact bushland (Chapter 5). While there are no specific areas 2850

where boobooks can be defined as verifiably extirpated, their density in the landscape 2851

appears to be substantially reduced in both of the two fragmented habitats, indicating lower 2852

habitat quality rather than the absence of usable space. Boobooks’ continued use of 2853

substantially altered landscapes is likely facilitated by the same traits which allow them to 2854

use the majority of vegetation types throughout Australia. 2855

The responses of particular species to habitat fragmentation can be impacted by 2856

species-specific traits including “trophic level, dispersal ability and degree of habitat 2857

specialisation” (Ewers and Didham, 2006). While traditional fragmentation models which 2858

assume a completely hostile matrix between islands of usable habitat are still likely to apply 2859

to species which are dietary or habitat specialists, they are probably less relevant when 2860

applied to more generalist species. Species responding to apparent fragmentation by 2861

making extensive use of resources in the matrix are often classified as “urban exploiters” 2862

(Conole and Kirkpatrick, 2011).The term “urban adapters” is often used to describe species 2863

with a lower capacity to tolerate urban development but these traits exist on a continuum 2864

(Callaghan et al., 2019). The same principle applies to species responding in a similar 2865

fashion to intensive agricultural land use. Boobooks continued presence in urban and 2866

agricultural landscapes, genetic connectivity, and observed capacity to use resources 2867

derived from urban and agricultural matrix coupled with reduced detection rates relative to 2868

areas of intact bushland suggest that they function as ‘adapters’ in urban and agricultural 2869

landscapes. Species which function as exploiters or adapters are probably inappropriate for 2870

use in modelling general impacts of habitat fragmentation because these groups 2871

disproportionally share generalist functional and morphological characteristics and are not 2872

representative of the previously existing suite of taxa present prior to extensive alteration of 2873

vegetation types (Conole and Kirkpatrick, 2011). 2874

Boobooks were initially chosen because they were present across all landscape types 2875

included in the study. Future studies specifically examining responses to habitat 2876

142

fragmentation should also consider key natural history traits linked to a species’ capacity to 2877

use altered habitat types when selecting study species. When dietary and habitat 2878

generalists with broad niche breadths are used, studies are less likely to detect impacts of 2879

threatening processes which may have already caused extirpation of even closely related 2880

species. For example, although boobooks in this study did not appear to be limited by nest 2881

hollow availability (Chapter 5), the congeneric Powerful Owl (Ninox strenua) appears to be 2882

limited due to its requirement of larger nest hollows which are less available in urban areas 2883

(Isaac et al., 2014a). 2884

If, on the other hand, a strong signal of detrimental effect from a specific proposed 2885

threatening process is apparent in species which are robust to overall habitat alteration, 2886

further evaluation of that threat over the spatial area where it is likely to occur is warranted. 2887

In the context of my study, the widespread and often severe exposure of boobooks to 2888

second generation anticoagulant rodenticides meets these criteria. As discussed in Lohr 2889

(2018), species of predatory birds with more specialised diets containing a larger proportion 2890

of rodents and species with larger home ranges are likely to be at greater risk of exposure to 2891

ARs. Anecdotally, raptor species meeting this description (i.e. Wedge-tailed Eagles (Aquila 2892

audax) and Masked Owls (Tyto novaehollandiae)) are largely or completely absent from the 2893

Perth metropolitan area but present outside its margins. Preliminary testing of these 2894

species (James Pay pers. comm., Michael Lohr unpublished data) has revealed exposure 2895

patterns in line with the predictions made in Lohr (2018). The detection of this pattern in 2896

boobooks despite their apparent capacity to use highly-altered landscapes, may be possible 2897

because the threat of SGARs itself is more recent than a large proportion of the habitat 2898

alteration as these chemicals were not invented until the late 1970s and early 1980s. 2899

Alternately, the continued presence of boobooks in urban areas despite the severity 2900

of the threat may be a function of their greater abundance on the wider landscape and 2901

immigration from adjacent unaffected areas masking the effects of suboptimal population 2902

parameters within urban areas. Under this circumstance, it is possible that the ability of 2903

boobooks to utilise highly-altered urban areas may pose a greater risk than if they actually 2904

experienced true habitat fragmentation with its concomitant loss of “usable space” and 2905

landscape permeability. Somewhat counterintuitively, a review of landscape-level 2906

fragmentation studies found documentation of positive effects of habitat fragmentation 2907

143

including “spreading of risk, reduced competition, and stabilization of predator-prey 2908

interactions” (Fahrig et al., 2019). The tendency of boobooks to use and capacity of 2909

boobooks to move through highly-altered habitats was demonstrated in Chapter 4. If 2910

demographic parameters in urban or agricultural landscapes are such that local populations 2911

are not numerically self-supporting and fragmentation does not pose an impediment to 2912

movement, these areas, at best, may be sinks where populations in altered habitats are 2913

subsidised by dispersal from areas of less-degraded habitat. 2914

If, however, maladaptive selection cues lead boobooks to preferentially select highly 2915

altered habitats, these areas may actually be ecological traps and could lead to declines in 2916

adjacent areas of objectively better quality habitat. A similar dynamic has been observed in 2917

Powerful Owls. Urban habitat does not appear to be an effective barrier to dispersal and 2918

fledglings have been observed travelling distances up to 18 km across urban habitat (Hogan 2919

and Cooke, 2010). In this species, habitat selection appears to be driven by availability of 2920

mammalian prey which are common in urban areas but a lack of available nesting hollows 2921

appears to create an ecological sink where breeding cannot occur (Isaac et al., 2014a). In 2922

our study of boobooks, nest site availability does not appear to be a limiting factor on 2923

abundance in urban and agricultural landscapes (Chapter 5). However, as demonstrated in 2924

Chapter 3, widespread and severe exposure to second generation anticoagulant 2925

rodenticides may reduce critical population parameters like survival and fecundity across 2926

age classes in areas of urban and periurban development. Determination of whether 2927

extensive conversion of bushland to urban or agricultural land uses creates an ecological 2928

sink or trap for boobooks would require quantification of population parameters including 2929

survival and fecundity in urban, agricultural, and bushland landscapes. However, 2930

degradation can likely be inferred from the dramatic differences in occupancy rates 2931

observed in Chapter 5. 2932

Care should be taken when interpreting the response of habitat generalists to novel 2933

anthropogenic land use types. Persistence in remnants surrounded by fragmentary matrix 2934

or even direct and consistent use of the matrix could indicate that the novel habitat type 2935

constitutes usable space of good or poor quality or it could indicate an ecological sink or 2936

trap. Distinguishing between these situations requires both knowledge of differences in 2937

demographic parameters between individuals using the fragmentary matrix and those using 2938

144

intact natural habitat, as well as an understanding of patterns of habitat selection relative to 2939

the availability of the habitats on a scale approximating Johnson’s second order selection 2940

(Johnson, 1980). Future attempts to understand landscape-level impacts of key threatening 2941

processes should incorporate observations of survival and fecundity as well as the use of 2942

GPS telemetry or other techniques which enable the observation of individual habitat 2943

selection in order to facilitate quantification of these parameters. 2944

Additionally, attempts to assess impacts of habitat fragmentation on wildlife need to 2945

consider how individual species respond to different types of matrix. While habitat 2946

fragmentation clearly exerts pressure on some wildlife populations through direct habitat 2947

loss and small population phenomena impacting remaining isolated populations, 2948

threatening processes flowing from specific matrix types also need to be considered in 2949

modelling impacts of habitat alteration. Species with large home ranges may be especially 2950

vulnerable to unconventional edge effects, particularly when the threats involve pathogens 2951

and pollutants which can have substantial impacts on exposed individuals even when a 2952

relatively small portion of their home range includes the land use type where the threat 2953

originates (Lohr, 2018). 2954

2955

Management Recommendations 2956

Anticoagulant Rodenticides 2957

Species which are resilient to habitat fragmentation can sometimes compensate for 2958

habitat loss by utilizing resources in the surrounding matrix (Ewers and Didham, 2006). In 2959

the case of boobooks, which are generalist predators (Higgins, 1999), this resource subsidy 2960

may come largely in the form of high abundances of introduced bird species and commensal 2961

rodents in the urban matrix. Anticoagulant rodenticide poisoning, especially by more 2962

persistent SGARs poses a serious threat to boobooks with home ranges containing urban 2963

and suburban habitat (Lohr, 2018). Subsequent testing of a larger suite of carnivorous 2964

wildlife in Australia – including species listed under the Australian Commonwealth 2965

Environment Protection and Biodiversity Conservation Act 1999 – has revealed similar 2966

patterns across the continent (Michael Lohr, unpublished data). The pervasive use of SGARs 2967

in and around areas of human habitation threatens to convert potential matrix subsidies 2968

145

into edge effects. Mitigating this threat will be important in maintaining boobooks and 2969

other carnivorous wildlife in urban systems and avoiding resultant tropic skew. In attempts 2970

to maintain native biodiversity in the face of extensive fragmentation the “loss of a few 2971

predator species often has impacts comparable in magnitude to those stemming from a 2972

large reduction in plant diversity” (Duffy, 2003). Bioaccumulation and biomagnification 2973

associated with highly-persistent SGARs make them particularly dangerous to non-target 2974

wildlife at the highest trophic levels and threaten to exacerbate trophic skew in already-2975

susceptible urban ecosystems, hastening ecosystem decay. 2976

Regulatory restrictions have been implemented in the United States and Canada (Lohr 2977

and Davis, 2018). Despite implementation of restrictions in the United States in 2011 2978

(Bradbury, 2008), a recent study indicated that mean exposure in raptors had not declined 2979

(Murray, 2017). Removal of SGARs from sale directly to the public is probably necessary but 2980

not sufficient to prevent severe and widespread exposure in urban and exurban carnivores. 2981

I recommend complete replacement of currently used SGARs with commercially available 2982

less-persistent alternatives including baits based on the FGARs warfarin and coumatetralyl, 2983

cholecalciferol, or corn gluten meal. This regulatory reform should be coupled with 2984

increased research into effective alternative solutions to rodent control problems to ensure 2985

maintenance of a suite of effective rodent control options which reduce the probability of 2986

secondary poisoning on non-target wildlife. 2987

Nest Box Supplementation 2988

Nest boxes are a popular conservation intervention particularly among community 2989

groups and have been promoted for use to aid boobooks (Hussey, 1997). Nest boxes are 2990

intended to increase availability of nesting hollows where their abundance has been 2991

reduced by loss or alteration of native vegetation. My research suggests that nest hollow 2992

availability is not likely to be a limiting factor for boobooks in the urban and agricultural 2993

remnant bushlands where they were tested. Nest site availability may be more limiting in 2994

areas of intensive human land use where remnant bushlands are absent, but consideration 2995

needs to be given as to whether the addition of nest boxes may incentivise use of areas that 2996

are otherwise unsuitable, creating the potential for an ecological trap. A better 2997

understanding of population parameters and natural hollow availability in such areas is 2998

needed before advocating large-scale use of nest boxes as a conservation measure for 2999

146

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

use. 3001

References 3002

Aguilar, R., Quesada, M., Ashworth, L., Herrerias-Diego, Y., Lobo, J., 2008. Genetic 3003

consequences of habitat fragmentation in plant populations: susceptible signals in 3004

plant traits and methodological approaches. Mol. Ecol. 17, 5177–5188. 3005

Albert, C.A., Wilson, L.K., Mineau, P., Trudeau, S., Elliott, J.E., 2010. Anticoagulant 3006

rodenticides in three owl species from Western Canada, 1988-2003. Arch. Environ. 3007

Contam. Toxicol. 58, 451–459. https://doi.org/10.1007/s00244-009-9402-z 3008

Alexander, W.B., 1921. The birds of the Swan River district, Western Australia. Emu 20, 149–3009

168. 3010

Almería, S., Calvete, C., Pagés, A., Gauss, C., Dubey, J.P., 2004. Factors affecting the 3011

seroprevalence of Toxoplasma gondii infection in wild rabbits (Oryctolagus cuniculus) 3012

from Spain. Vet. Parasitol. 123, 265–270. https://doi.org/10.1016/j.vetpar.2004.06.010 3013

Alterio, N., 1996. Secondary poisoning of stoats (Mustela erminea), feral ferrets (Mustela 3014

furo), and feral house cats (Felis catus) by the anticoagulant poison, brodifacoum. New 3015

Zeal. J. Zool. 23, 331–338. https://doi.org/10.1080/03014223.1996.9518092 3016

Andrén, H., 1994. Effects of Habitat Fragmentation on Birds and Mammals in Landscapes 3017

with Different Proportions of Suitable Habitat: A Review. Oikos 71, 355–366. 3018

https://doi.org/10.2307/3545823 3019

Arling, T., Yolken, R., Lapidus, M., Langenberg, P., Dickerson, F.B., Zimmerman, S.A., Balis, T., 3020

Cabassa, J.A., Scrandis, D.A., Tonelli, L.H., Postolache, T.T., 2009. Toxoplasma gondii 3021

antibody titers and history of suicide attempts in patients with recurrent mood 3022

disorders. J. Nerv. Ment. Dis. 3, 905–908. 3023

Aubert, D., Terrier, M.-E., Dumètre, A., Barrat, J., Villena, I., 2008. Prevalence of Toxoplasma 3024

gondii in raptors from France. J. Wildl. Dis. 44, 172–173. https://doi.org/10.7589/0090-3025

3558-44.1.172 3026

Aumann, T., 2001. An intraspecific and interspecific comparison of raptor diets in the south-3027

147

west of the Northern Territory, Australia. Wildl. Res. 28, 379–393. 3028

Australian Government Department of Health: Therapeutic Goods Administration, 2017. 3029

The Poisons Standard June 2017. Canberra. 3030

Australian Pesticides and Veterinary Medicines Authority, 2017a. Commonwealth of 3031

Australia Gazette: Agricultural and Veterinary Chemicals 6, 18. 3032

Australian Pesticides and Veterinary Medicines Authority, 2017b. Agricultural And 3033

Veterinary Permits Search [WWW Document]. URL 3034

https://portal.apvma.gov.au/permits?p_auth=MpDv1rrW&p_p_id=permitsportlet_WA3035

R_permitsportlet&p_p_lifecycle=1&p_p_state=normal&p_p_mode=view&p_p_col_id=3036

column-3037

1&p_p_col_pos=1&p_p_col_count=3&_permitsportlet_WAR_permitsportlet_javax.por3038

tlet.action=search (accessed 8.4.17). 3039

Australian Pesticides and Veterinary Medicines Authority, 2015. Second generation anti-3040

coagulant rodenticides—priority 2 [WWW Document]. URL 3041

https://apvma.gov.au/node/19286 (accessed 7.14.17). 3042

Barrett, G., Silcocks, A., Barry, S., Cunningham, R., Poulter, R., 2003. The New Atlas of 3043

Australian Birds. Royal Ornithologists Union, Hawthorn East, Victoria. 3044

Barros, M., Cabezón, O., Dubey, J.P., Almería, S., Ribas, M.P., Escobar, L.E., Ramos, B., 3045

Medina-Vogel, G., 2018. Toxoplasma gondii infection in wild mustelids and cats across 3046

an urban-rural gradient. PLoS One 13, 1–16. 3047

https://doi.org/10.1371/journal.pone.0199085 3048

Battin, J., 2004. When good animals love bad habitats: Ecological traps and the conservation 3049

of animal populations. Conserv. Biol. 18, 1482–1491. https://doi.org/10.1111/j.1523-3050

1739.2004.00417.x 3051

Beckingham, D., 2012. Birders Spot an Amazing 47 Species. Friends Lake Claremont News 6–3052

7. 3053

Belo, N.O., Pinheiro, R.T., Reis, E.S., Ricklefs, R.E., Braga, É.M., 2011. Prevalence and lineage 3054

diversity of avian haemosporidians from three distinct cerrado habitats in Brazil. PLoS 3055

148

One 6, 1–9. https://doi.org/10.1371/journal.pone.0017654 3056

Bender, D.J., Fahrig, L., 2005. Matrix structure obscures the relationship between interpatch 3057

movement and patch size and isolation. Ecology 86, 1023–1033. 3058

https://doi.org/10.1890/03-0769 3059

Bennison, C., Friend, J.A., Button, T., Mills, H., Lambert, C., Bencini, R., 2016. Potential 3060

impacts of poison baiting for introduced house mice on native animals on islands in 3061

Jurien Bay , Western Australia. Wildl. Res. 43, 61–68. 3062

Bettink, K., 2015. Control and eradication of black rats (*Rattus rattus) on Penguin Island, 3063

Western Australia, December 2012 – December 2014. Perth, Western Australia. 3064

Beyer, G.L., Goldingay, R.L., 2006. The value of nest boxes in the research and management 3065

of Australian hollow-using arboreal marsupials. Wildl. Res. 33, 161–174. 3066

https://doi.org/10.1071/WR04109 3067

BirdLife Australia, 2015. The State of Australia’s Birds 2015. Melbourne. 3068

Bishop, C.A., Williams, K.E., Kirk, D.A., Nantel, P., Reed, E., Elliott, J.E., Kirk, D.A., Reed, E., 3069

Elliott, J.E., 2016. A population model of the impact of a rodenticide containing 3070

strychnine on Great Basin Gophersnakes (Pituophis catenifer deserticola). 3071

Ecotoxicology 25, 1390–1405. https://doi.org/10.1007/s10646-016-1690-2 3072

Block, W.M., Brennan, L.A., 1993. The habitat concept in ornithology: Theory and 3073

applications, in: Power, D.M. (Ed.), Current Ornithology. New York, pp. 35–91. 3074

https://doi.org/10.1007/978-1-4757-9912-5_2 3075

Boal, C.W., Mannan, R.W., 1999. Comparative breeding ecology of Cooper’s hawks in urban 3076

and exurban areas of southeastern Arizona. J. Wildl. Manage. 63, 77–84. 3077

https://doi.org/10.2307/3802488 3078

Bohonak, A.J., 1999. Dispersal, gene flow, and population structure. Q. Rev. Biol. 74, 21–45. 3079

https://doi.org/10.1086/392950 3080

Bortolotti, G.R., 1994. Effect of nest-box size on nest-site preference and reproduction in 3081

American Kestrels. J. Raptor Res. 28, 127–133. 3082

Bradbury, S., 2008. Risk Mitigation Decision for Ten Rodenticides. Washington D. C., USA. 3083

149

Brakes, C.R., Smith, R.H., 2005. Exposure of non-target small mammals to rodenticides : 3084

Short-term effects , recovery and implications for secondary Poisoning. J. Appl. Ecol. 3085

42, 118–128. https://doi.org/10.1111/j.1365-2664.2005.00997.x 3086

Brash, A.R., 1987. The history of avian extinction and forest conversion on Puerto Rico. Biol. 3087

Conserv. https://doi.org/10.1016/0006-3207(87)90028-0 3088

Brooker, M.G., Ridpath, M.G., 1980. The Diet of the Wedge-Tailed Eagle, Aquila audax, in 3089

Western Australia. Aust. Wildl. Res. 7, 433–452. https://doi.org/10.1071/WR9800433 3090

Brown, P.R., Singleton, G.R., 1998. Efficacy of brodifacoum to control house mice, Mus 3091

domesticus, in wheat crops in Southern Australia. Crop Prot. 17, 345–352. 3092

https://doi.org/10.1016/S0261-2194(98)00026-X 3093

Burbidge, A.A., 2004. Montebello Renewal: Western Shield review — February 2003. 3094

Conserv. Sci. West. Aust. 5, 194–201. 3095

Burbidge, A.A., Abbott, I., Comer, S., Adams, E., Berry, O., Penwarden, K.E., 2012. 3096

Unforeseen consequences of a misidentified rodent: case study from the Archipelago 3097

of the Recherche, Western Australia. Aust. Mammal. 34, 55–58. 3098

https://doi.org/http://dx.doi.org/10.1071/AM11004 3099

Burbidge, A.A., McKenzie, N.L., 1989. Patterns in the Modern Decline of Western Australia’ s 3100

Vertebrate Fauna: Causes and Conservation Implications. Biol. Conserv. 50, 143–198. 3101

Burbidge, A.A., Morris, K.D., 2002. Introduced mammal eradications for nature conservation 3102

on Western Australian islands : a review, in: Veitch, C.R., Clout, M.N., Towns, D.R. 3103

(Eds.), Turning the Tide: The Eradication of Invasive Species. IUCN SSC Invasive Species 3104

Specialist Group, Gland, Switzerland, pp. 64–70. 3105

Burghardt, K.T., Tallamy, D.W., Shriver, W.G., 2009. Impact of native plants on bird and 3106

butterfly biodiversity in suburban landscapes. Conserv. Biol. 23, 219–224. 3107

https://doi.org/10.1111/j.1523-1739.2008.01076.x 3108

Butler, S., 2010. The sub-lethal effects of Second Generation Anticoagulant Rodenticides on 3109

birds Thesis submitted for the degree of Doctor of Philosophy at the University of 3110

Leicester by Sarah Elizabeth Butler Department of Biology University of Leicester 3111

150

September 2010 T. 3112

Cabezón, O., García-Bocanegra, I., Molina-López, R., Marco, I., Blanco, J.M., Höfle, U., 3113

Margalida, A., Bach-Raich, E., Darwich, L., Echeverría, I., Obón, E., Hernández, M., Lavín, 3114

S., Dubey, J.P., Almería, S., 2011. Seropositivity and risk factors associated with 3115

Toxoplasma gondii infection in wild birds from Spain. PLoS One 6, 1–7. 3116

https://doi.org/10.1371/journal.pone.0029549 3117

Callaghan, C.T., Major, R.E., Wilshire, J.H., Martin, J.M., Kingsford, R.T., Cornwell, W.K., 2019. 3118

Generalists are the most urban-tolerant of birds: a phylogenetically controlled analysis 3119

of ecological and life history traits using a novel continuous measure of bird responses 3120

to urbanization. Oikos 1–14. https://doi.org/10.1111/oik.06158 3121

Campbell, K.J., Beek, J., Eason, C.T., Glen, A.S., Godwin, J., Gould, F., Holmes, N.D., Howald, 3122

G.R., Madden, F.M., Ponder, J.B., Threadgill, D.W., Wegmann, A.S., Baxter, G.S., 2015. 3123

The next generation of rodent eradications: Innovative technologies and tools to 3124

improve species specificity and increase their feasibility on islands. Biol. Conserv. 185, 3125

47–58. https://doi.org/10.1016/j.biocon.2014.10.016 3126

Canfield, P.J., Hartley, W.J., Dubey, J.P., 1990. Lesions of toxoplasmosis in Australian 3127

marsupials. J. Comp. Pathol. 103, 159–167. https://doi.org/10.1016/S0021-3128

9975(08)80172-7 3129

Cardador, L., Carrete, M., Mañosa, S., 2011. Can intensive agricultural landscapes favour 3130

some raptor species? The Marsh harrier in north-eastern Spain. Anim. Conserv. 14, 3131

382–390. https://doi.org/10.1111/j.1469-1795.2011.00449.x 3132

Carrete, M., Tella, J., Blanco, G., Bertellotti, M., 2009. Effects of habitat degradation on the 3133

abundance, richness and diversity of raptors across Neotropical biomes. Biol. Conserv. 3134

142, 2002–2011. 3135

Chen, J.C., Tsai, Y.J., Wu, Y.L., 2015. Seroprevalence of Toxoplasma gondii antibodies in wild 3136

birds in Taiwan. Res. Vet. Sci. 102, 184–188. https://doi.org/10.1016/j.rvsc.2015.08.010 3137

Cherriman, S.C., 2007. Territory size and diet throughout the year of the Wedge-tailed Eagle 3138

(Aquila audax) in the Perth region, Western Australia. Curtin University. 3139

151

Choquenot, D., Barry, K., Lukins, B., 1990. An evaluation of warfarin for the control of feral 3140

pigs. J. Wildl. Manage. 54, 353–359. 3141

Christensen, T.K., Lassen, P., Elmeros, M., 2012. High Exposure Rates of Anticoagulant 3142

Rodenticides in Predatory Bird Species in Intensively Managed Landscapes in Denmark. 3143

Arch. Environ. Contam. Toxicol. 63, 437–444. https://doi.org/10.1007/s00244-012-3144

9771-6 3145

Collinge, S.K., 1996. Ecological consequences of habitat fragmentation: Implications for 3146

landscape architecture and planning. Landsc. Urban Plan. 36, 59–77. 3147

https://doi.org/10.1016/S0169-2046(96)00341-6 3148

Commonwealth of Australia, 2015. Australian Bird and Bat Banding Scheme website [WWW 3149

Document]. URL http://www.environment.gov.au/cgi-bin/biodiversity/abbbs/abbbs-3150

search.pl?taxon_id=81095 (accessed 4.29.15). 3151

Commonwealth of Australia, 1994. Agricultural and Veterinary Chemicals Code Act 1994. 3152

Australia. 3153

Conole, L.E., Kirkpatrick, J.B., 2011. Functional and spatial differentiation of urban bird 3154

assemblages at the landscape scale. Landsc. Urban Plan. 100, 11–23. 3155

https://doi.org/10.1016/j.landurbplan.2010.11.007 3156

Cooke, R., Wallis, R., Hogan, F., White, J., Webster, A., 2006. The diet of powerful owls 3157

(Ninox strenua) and prey availability in a continuum of habitats from disturbed urban 3158

fringe to protected forest environments in south-eastern Australia. Wildl. Res. 33, 199–3159

206. https://doi.org/10.1071/WR05058 3160

Cory, F., Wilson, A., Priddel, D., Carlile, N., Klomp, N., 2011. Eradication of the House Mouse 3161

(Mus musculus) from Montague Island, New South Wales, Australia. Ecol. Manag. 3162

Restor. 12, 102–109. https://doi.org/10.1111/j.1442-8903.2011.00583.x 3163

Cowled, B.D., Elsworth, P., Lapidge, S.J., 2008. Additional toxins for feral pig (Sus scrofa) 3164

control: Identifying and testing Achilles’ heels. Wildl. Res. 35, 651–662. 3165

https://doi.org/10.1071/WR07072 3166

Cox, P., Smith, R.H., 1992. Rodenticide ecotoxicology: pre-lethal effects of anticoagulants on 3167

152

rat behaviour, in: Vertebrate Pest Conference Proceedings Collection. Lincoln, 3168

Nebraska, pp. 165–170. 3169

Cypher, B.L., McMillin, S.C., Westall, T.L., Job, C.V.H., Hosea, R.C., Finlayson, B.J., 2014. 3170

Rodenticide Exposure Among Endangered Kit Foxes Relative to Habitat Use in an Urban 3171

Landscape. Cities Environ. 7, Article 8. 3172

Czechura, G.V., Debus, S.J.S., Mooney, N.J., 1987. The Collared Sparrowhawk Accipiter 3173

cirrhocephalus: a review and comparison with the Brown Goshawk Accipiter fasciatus. 3174

Aust. Bird Watch. 12, 35–62. 3175

Davies-Colley, R.J., Payne, G.W., van Elswijk, M., 2000. Microclimate gradient across a forest 3176

edge. N. Z. J. Ecol. 24, 111–121. 3177

Davis, A., Major, R.E.R., Taylor, C.E.C., 2014. Distribution of tree-hollows and hollow 3178

preferences by parrots in an urban landscape. Emu 114, 295–303. 3179

https://doi.org/http://dx.doi.org/10.1071/MU13065 3180

Davis, R.A., Wilcox, J.A., 2013. Adapting to suburbia: Bird ecology on an urbanbushland 3181

interface in Perth, Western Australia. Pacific Conserv. Biol. 19, 110–120. 3182

https://doi.org/10.1071/PC130110 3183

Debus, S.J., 2009. The owls of Australia: a field guide to Australian night birds. Envirobook, 3184

Canterbury, NSW. 3185

Debus, S.J.S., 2012. Norfolk Island Boobook chick deaths. Boobook 30, 6. 3186

Delgado-V., C.A., French, K., 2012. Parasite–bird interactions in urban areas: Current 3187

evidence and emerging questions. Landsc. Urban Plan. 105, 5–14. 3188

https://doi.org/10.1016/j.landurbplan.2011.12.019 3189

Department of Agriculture and Food Western Australia, 2015. Rabbit control: Bait products 3190

[WWW Document]. URL https://www.agric.wa.gov.au/1080/rabbit-control-bait-3191

products?page=0,2 (accessed 6.15.15). 3192

Didham, R.K., Lawton, J.H., Hammond, P.M., Eggleton, P., 1998. Trophic structure stability 3193

and extinction dynamics of beetles (Coleoptera) in tropical forest fragments. Philos. 3194

Trans. R. Soc. B-Biological Sci. 353, 437–451. https://doi.org/10.1098/rstb.1998.0221 3195

153

Doherty, T.S., Davis, R.A., van Etten, E.J.B., Algar, D., Collier, N., Dickman, C.R., Edwards, G., 3196

Masters, P., Palmer, R., Robinson, S., 2015. A continental-scale analysis of feral cat diet 3197

in Australia. J. Biogeogr. 42, 964–975. https://doi.org/10.1111/jbi.12469 3198

Donald, P.F., Green, R.E., Heath, M.F., Donal, P.F., Gree, R.E., Heath, M.F., 2001. Agricultural 3199

intensification and the collapse of Europe’s farmland bird populations. Proc. R. Soc. 3200

London Ser. B-Biological Sci. 268, 25–29. https://doi.org/10.1098/rspb.2000.1325 3201

Dowding, J.E., Lovegrove, T. im G., Kast, S.N., Puckett, M., 2006. Mortality of northern New 3202

Zealand dotterels (Charadrius obscurus aquilonius) following an aerial poisoning 3203

operation. Notornis 53, 235–239. 3204

Dowding, J.E., Murphy, E.C., Veitch, C.R., 1999. Brodifacoum residues in target an non-target 3205

species following an aerial poisoning operation on Motuihe Island, Hauraki Gulf, New 3206

Zealand. N. Z. J. Ecol. 23, 207–214. 3207

Dowding, C. V., Shore, R.F., Worgan, A., Baker, P.J., Harris, S., 2010. Accumulation of 3208

Anticoagulant Rodenticides in a Non-target Insectivore, the European hedgehog 3209

(Erinaceus europaeus). Environamental Pollut. 158, 161–166. 3210

Dubey, J.P., 2010. Toxoplasma gondii infections in chickens (Gallus domesticus): prevalence, 3211

clinical disease, diagnosis and public health significance. Zoonoses Public Health 57, 3212

60–73. https://doi.org/10.1111/j.1863-2378.2009.01274.x 3213

Dubey, J.P., 2002. A review of toxoplasmosis in wild birds. Vet. Parasitol. 106, 121–153. 3214

https://doi.org/10.1016/S0304-4017(02)00034-1 3215

Dubey, J.P., Felix, T. a, Kwok, O.C.H., 2010. Serological and parasitological prevalence of 3216

Toxoplasma gondii in wild birds from Colorado. J. Parasitol. 96, 937–939. 3217

https://doi.org/10.1645/GE-2501.1 3218

Dubey, J.P., Jones, J.L., 2008. Toxoplasma gondii infection in humans and animals in the 3219

United States. Int. J. Parasitol. 38, 1257–78. 3220

https://doi.org/10.1016/j.ijpara.2008.03.007 3221

Dubey, J.P., Porter, S.L., Tseng, F., Shen, S.K., Thulliez, P., 1992. Induced Toxoplasmosis in 3222

Owls. J. Zoo Wildl. Med. 23, 98–102. 3223

154

Duckett, J.E., Karuppiah, S., 1990. A guide to the planter in utilizing barn owls (Tyto alba) as 3224

an effective biological control in mature oil palm plantations, in: Jalani, S., Zin, Z.Z., 3225

Paranjothy, K., Ariffin, D., Rajanaidu, N., Cheah, S.C., Basri, M.W., Henson, I. (Eds.), 3226

Proceedings of the 1989 International Palm Oil Development Conference-Agriculture. 3227

Palm Oil Research Institute of Malaysia, pp. 357–372. 3228

Duffy, J.E., 2003. Biodiversity loss, trophic skew and ecosystem functioning. Ecol. Lett. 6, 3229

680–687. https://doi.org/10.1046/j.1461-0248.2003.00494.x 3230

Dunlop, J.N., Rippey, E., Bradshaw, L.E., Burbidge, A.A., 2015. Recovery of seabird colonies 3231

on Rat Island (Houtman Abrolhos) following the eradication of introduced predators. J. 3232

R. Soc. West. Aust. 98, 29–36. 3233

Durant, R., 2006. Identifying factors that influence the occupancy of nest boxes by arboreal 3234

mammals in peri-urban landscapes. Charles Sturt University. 3235

Durant, R., Luck, G.W., Matthews, A., 2009. Nest-box use by arboreal mammals in a peri-3236

urban landscape. Wildl. Res. 36, 565–573. https://doi.org/10.1071/WR09058 3237

Earl, D.A., vonHoldt, B.M., 2012. STRUCTURE HARVESTER: A website and program for 3238

visualizing STRUCTURE output and implementing the Evanno method. Conserv. Genet. 3239

Resour. 4, 359–361. https://doi.org/10.1007/s12686-011-9548-7 3240

Eason, C.T., Wright, G.R., Batcheler, D., 1996. Anticoagulant effects and the persistence of 3241

brodifacoum in possums (Trichosurus vulpecula). New Zeal. J. Agric. Res. 39, 397–400. 3242

https://doi.org/10.1080/00288233.1996.9513199 3243

Eason, C.T., Wright, G.R.G., Milne, L.M., Morriss, G.A., 2001. Laboratory and field studies of 3244

brodifacoum residues in relation to risk of exposure to wildlife and people. Sci. 3245

Conserv. 177, 11–23. 3246

Eisemann, J.D., Swift, C.E., 2006. Ecological and Human Health Hazards from Broadcast 3247

Application of 0.005% Diphacinone Rodenticide Baits in Native Hawaiian Ecosystems, 3248

in: Timm, R.M., O’Brien, J.M. (Eds.), Proceedings of the Vertebrate Pest Conference. 3249

University of California, Davis, California, pp. 413–433. 3250

El-Massry, A., Mahdy, O.A., El-Ghaysh, A., Dubey, J.P., 2000. Prevalence of Toxoplasma 3251

155

gondii Antibodies in Sera of Turkeys, Chickens, and Ducks from Egypt. J. Parasitol. 86, 3252

627–628. https://doi.org/10.1645/0022-3395(2000)086[0627:POTGAI]2.0.CO;2 3253

Elliott, J.E., Hindmarch, S., Albert, C.A., Emery, J., Mineau, P., Maisonneuve, F., 2014. 3254

Exposure pathways of anticoagulant rodenticides to nontarget wildlife. Environ. Monit. 3255

Assess. 186, 895–906. https://doi.org/10.1007/s10661-013-3422-x 3256

Erickson, W., Urban, D., 2004. Potential Risk of Nine Rodenticides to Birds and Mammals: A 3257

Comparative Approach. Washington DC. 3258

Evanno, G., Regnaut, S., Goudet, J., 2005. Detecting the number of clusters of individuals 3259

using the software STRUCTURE: A simulation study. Mol. Ecol. 14, 2611–2620. 3260

https://doi.org/10.1111/j.1365-294X.2005.02553.x 3261

Ewers, R.M., Didham, R.K., 2006. Confounding factors in the detection of species responses 3262

to habitat fragmentation. Biol. Rev. 81, 117–142. 3263

https://doi.org/10.1017/S1464793105006949 3264

Fahrig, L., 1997. Relative effects of habitat loss and fragmentation on population extinction. 3265

J. Wildl. Manage. 61, 603–610. 3266

Fahrig, L., Arroyo-Rodríguez, V., Bennett, J.R., Boucher-Lalonde, V., Cazetta, E., Currie, D.J., 3267

Eigenbrod, F., Ford, A.T., Harrison, S.P., Jaeger, J.A.G., Koper, N., Martin, A.E., Martin, 3268

J.-L., Metzger, J.P., Morrison, P., Rhodes, J.R., Saunders, D.A., Simberloff, D., Smith, 3269

A.C., Tischendorf, L., Vellend, M., Watling, J.I., 2019. Is habitat fragmentation bad for 3270

biodiversity? Biol. Conserv. 230, 179–186. 3271

https://doi.org/10.1016/j.biocon.2018.12.026 3272

Fancourt, B.A., 2016. Diagnosing species decline : a contextual review of threats , causes and 3273

future directions for management and conservation of the eastern quoll. Wildl. Res. 43, 3274

197–211. 3275

Fancourt, B.A., 2014. Rapid decline in detections of the Tasmanian bettong (Bettongia 3276

gaimardi) following local incursion of feral cats (Felis catus). Aust. Mammal. 36, 247–3277

253. https://doi.org/10.1071/AM14004 3278

Fiore, C.E., Tamburino, C., Foti, R., Grimaldi, D., 1990. Reduced axial bone mineral content in 3279

156

patients taking an oral anticoagulant. South. Med. J. 83, 538–542. 3280

https://doi.org/10.1097/00007611-199005000-00013 3281

Fisher, P., Brown, S., Arrow, J., 2015. Pindone residues in rabbit tissues : implications for 3282

secondary hazard and risk to non-target wildlife. Wildl. Res. 42, 362–370. 3283

Fisher, P., O’Connor, C.., Wright, G., Eason, C.T., 2003. Persistence of four anticoagulant 3284

rodenticides in the livers of laboratory rats, DOC Science Internal Series. Wellington, 3285

New Zealand. 3286

Flegr, J., Havlícek, J., Kodym, P., Malý, M., Smahel, Z., 2002. Increased risk of traffic accidents 3287

in subjects with latent toxoplasmosis: a retrospective case-control study. BMC Infect. 3288

Dis. 2, 1–6. https://doi.org/10.1186/1471-2334-2-11 3289

Flegr, J., Prandota, J., Sovičková, M., Israili, Z.H., 2014. Toxoplasmosis--a global threat. 3290

Correlation of latent toxoplasmosis with specific disease burden in a set of 88 3291

countries. PLoS One 9, e90203. https://doi.org/10.1371/journal.pone.0090203 3292

Foley, J., Clifford, D., Castle, K., Cryan, P., Ostfeld, R.S., 2011. Investigating and managing the 3293

rapid emergence of white-nose syndrome, a novel, fatal, infectious disease of 3294

hibernating bats. Conserv. Biol. 25, 223–31. https://doi.org/10.1111/j.1523-3295

1739.2010.01638.x 3296

Ford, H. a, Barrett, G.W., Saunders, D.A., Recher, H.F., 2001. Why have birds in the 3297

woodland of Southern Australia declined? Biol. Conserv. 97, 71–88. 3298

Fourel, I., Sage, M., Benoit, E., Lattard, V., 2018. Liver and fecal samples suggest differential 3299

exposure of red fox (Vulpes vulpes) to trans- and cis-bromadiolone in areas from France 3300

treated with plant protection products. Sci. Total Environ. 622–623, 924–929. 3301

https://doi.org/10.1016/j.scitotenv.2017.12.053 3302

Fox, J.C., Hamilton, F., Occhipinti, S., 2009. Tree hollow incidence in Victorian state forests. 3303

Aust. For. 72, 39–48. https://doi.org/10.1080/00049158.2009.10676288 3304

Franklin, A.B., Carlson, P.C., Rex, A., Rockweit, J.T., Garza, D., Culhane, E., Volker, S.F., Dusek, 3305

R.J., Shearn-Bochsler, V.I., Gabriel, M.W., Horak, K.E., 2018. Grass is not always 3306

greener: Rodenticide exposure of a threatened species near marijuana growing 3307

157

operations. BMC Res. Notes 11, 1–7. https://doi.org/10.1186/s13104-018-3206-z 3308

Franklin, A.B., Noon, B.R., George, T.L., 2002. What is habitat fragmentation? Stud. Avian 3309

Biol. 25, 20–29. 3310

Franks, A., Franks, S., 2006. Nest Boxes for Wildlife: A Practical Guide. Blooming Books, 3311

Melbourne, Australia. 3312

Freeland, W.J., 1994. Parasites, pathogens and the impacts of introduced organisms on the 3313

balance of nature in Australia, in: Moritz, C., Kikkawa, J. (Eds.), Conservation Biology in 3314

Australia and Oceania. Chipping Norton, N.S.W., pp. 171–180. 3315

Freeman, A.B., Hickling, J.G., Bannock, C.A., 1996. Response of the Skink Oligosoma 3316

maccanni (Reptilia : Lacertilia) to Two Vertebrate Pest- Control Baits. Wildl. Res. 23, 3317

511–516. https://doi.org/10.1071/WR9960511 3318

Froeschke, G., van der Mescht, L., McGeoch, M., Matthee, S., 2013. Life history strategy 3319

influences parasite responses to habitat fragmentation. Int. J. Parasitol. 43, 1109–1118. 3320

https://doi.org/10.1016/j.ijpara.2013.07.003 3321

Fuentes, E., Olsen, J., Rose, A.B., 2005. Breeding Diet at two Whistling Kite nests near 3322

Canberra. Aust. F. Ornithol. 22, 122–125. 3323

Gabriel, M.W., Woods, L.W., Poppenga, R., Sweitzer, R.A., Thompson, C., Matthews, S.M., 3324

Higley, J.M., Keller, S.M., Purcell, K., Barrett, R.H., Wengert, G.M., Sacks, B.N., Clifford, 3325

D.A.L., 2012. Anticoagulant rodenticides on our public and community lands: Spatial 3326

distribution of exposure and poisoning of a rare forest carnivore. PLoS One 7, e40163. 3327

https://doi.org/10.1371/journal.pone.0040163 3328

Gazzonis, A.L., Zanzani, S.A., Santoro, A., Veronesi, F., Olivieri, E., Villa, L., Lubian, E., Lovati, 3329

S., Bottura, F., Epis, S., Manfredi, M.T., 2018. Toxoplasma gondii infection in raptors 3330

from Italy: Seroepidemiology and risk factors analysis. Comp. Immunol. Microbiol. 3331

Infect. Dis. 60, 42–45. https://doi.org/10.1016/j.cimid.2018.10.002 3332

Gehlbach, F., 1996. A Model of Raptor Urbanization, in: Bird, D.M., Varland, D.E., Negro, J.J. 3333

(Eds.), Raptors in Human Landscapes: Adaptations to Built and Cultivated 3334

Environments. Academic Press, New York, pp. 69–74. 3335

158

Gibbons, P., Lindenmayer, D.B., Barry, S.C., Tanton, M.T., 2002. Hollow selection by 3336

vertebrate fauna in forests of southeastern Australia and implications for forest 3337

management. Biol. Conserv. 103, 1–12. https://doi.org/10.1016/S0006-3207(01)00109-3338

4 3339

Gibbons, P., Lindenmayer, D.B., Barry, S.C., Tanton, M.T., 2000. Hollow formation in 3340

eucalypts from temperate forests in Southeastern Australia. Pacific Conserv. Biol. 6, 3341

218–228. https://doi.org/10.1071/PC000217 3342

Gilbert, F., Gonzalez, A., Evans-Freke, I., 1998. Corridors maintain species richness in the 3343

fragmented landscapes of a microecosystem. Proc. R. Soc. London. Ser. B Biol. Sci. 265, 3344

577–582. 3345

Godfrey, M.E.R., 1984. Acute toxicity of brodifacoum to wallabies (Macropus rufogriseus). 3346

New Zeal. J. Exp. Agric. 12, 63–64. https://doi.org/10.1080/03015521.1984.10427791 3347

Goldingay, R.L., Stevens, J.R., 2009. Use of artificial tree hollows by Australian birds and 3348

bats. Wildl. Res. 36, 81–97. https://doi.org/10.1071/WR08064 3349

Gracey, M., 2000. Historical, cultural, political, and social influences on dietary patterns and 3350

nutrition in Australian Aboriginal children 1 – 4. Am. J. Clin. Nutrtion 72, 1361–1367. 3351

Grarock, K., Tidemann, C.R., Wood, J., Lindenmayer, D.B., 2012. Is it benign or is it a Pariah? 3352

Empirical evidence for the impact of the common Myna (Acridotheres tristis) on 3353

Australian birds. PLoS One 7, e40622. https://doi.org/10.1371/journal.pone.0040622 3354

Guichoux, E., Lagache, L., Wagner, S., Chaumeil, P., Léger, P., Lepais, O., Lepoittevin, C., 3355

Malausa, T., Revardel, E., Salin, F., Petit, R.J., 2011. Current trends in microsatellite 3356

genotyping. Mol. Ecol. Resour. 11, 591–611. https://doi.org/10.1111/j.1755-3357

0998.2011.03014.x 3358

Guthery, F.S., Rybak, A.R., Walsh, W.R., Fuhlendorf, S.D., Hiller, T.I.M.L., 2005. Quantifying 3359

usable space for wildlife with use-availability data. J. Wildl. Manage. 69, 655–663. 3360

Gwee, C.Y., Christidis, L., Eaton, J.A., Norman, J.A., Trainor, C.R., Verbelen, P., Rheindt, F.E., 3361

2017. Bioacoustic and multi-locus DNA data of Ninox owls support high incidence of 3362

extinction and recolonisation on small, low-lying islands across Wallacea. Mol. 3363

159

Phylogenet. Evol. 109, 246–258. https://doi.org/10.1016/j.ympev.2016.12.024 3364

Hager, S.B., 2009. Human-Related Threats to Urban Raptors. J. Raptor Res. 43, 210–226. 3365

https://doi.org/10.3356/JRR-08-63.1 3366

Haig, S.M., Steven Wagner, R., Forsman, E.D., Mullins, T.D., 2001. Geographic variation and 3367

genetic structure in Spotted Owls. Conserv. Genet. 2, 25–40. 3368

https://doi.org/10.1023/A:1011561101460 3369

Haila, Y., 2002. A Conceptual Genealogy of Fragmentation Research : From Island 3370

Biogeography to Landscape Ecology. Ecol. Appl. 12, 321–334. 3371

Hall, L.S., Krausman, P.R., Morrison, M.L., 1997. The habitat concept and a plea for standard 3372

terminology. Wildl. Soc. Bull. 25, 173–182. 3373

Hansen, M.J., Clevenger, A.P., 2005. The influence of disturbance and habitat on the 3374

presence of non-native plant species along transport corridors. Biol. Conserv. 125, 249–3375

259. https://doi.org/10.1016/j.biocon.2005.03.024 3376

Harper, M.J., McCarthy, M.A., Van Der Ree, R., 2005. The abundance of hollow-bearing trees 3377

in urban dry sclerophyll forest and the effect of wind on hollow development. Biol. 3378

Conserv. 122, 181–192. https://doi.org/10.1016/j.biocon.2004.07.003 3379

Harrison, S., Taylor, A.D., 1997. Empirical evidence for metapopulation dynamics, in: Hanski, 3380

I.A., Gilpin, M.E. (Eds.), Metapopulation Biology: Ecology, Genetics, and Evolution. 3381

Academic Press, San Diego, California, USA, pp. 27–42. 3382

Hartley, W.J., Dubey, J.P., 1991. Fatal toxoplasmosis in some native Australian birds. J. Vet. 3383

Diagnostic Investig. 3, 167–169. https://doi.org/10.1177/104063879100300213 3384

Havlícek, J., Gasová, Z.G., Smith, A.P., Zvára, K., Flegr, J., 2001. Decrease of psychomotor 3385

performance in subjects with latent “asymptomatic” toxoplasmosis. Parasitology 122, 3386

515–520. https://doi.org/10.1017/S0031182001007624 3387

Health Canada: Pest Management Regulatory Agency, 2010. Risk Mitigation Measures for 3388

Eight Rodenticides. 3389

Hejlícek, K., Literák, I., Nezval, J., 1997. Toxoplasmosis in wild mammals from the Czech 3390

Republic. J. Wildl. Dis. 33, 480–485. 3391

160

Herring, G., Eagles-Smith, C.A., Buck, J., 2017. Characterizing Golden Eagle Risk to Lead and 3392

Anticoagulant Rodenticide Exposure: A Review. J. Raptor Res. 51, 273–292. 3393

Higgins, P.J. (Ed.), 1999. Southern Boobook, in: Handbook of Australian, New Zealand and 3394

Antarctic Birds. Volume 4: Parrots to Dollarbird. Oxford University Press, Melbourne, 3395

pp. 864–875. 3396

Hindmarch, S., Elliott, J.E., 2014. A specialist in the city: the diet of barn owls along a rural to 3397

urban gradient. Urban Ecosyst. 18, 477–488. https://doi.org/10.1007/s11252-014-3398

0411-y 3399

Hoare, J.M., Hare, K.M., 2006. The impact of brodifacoum on non-target wildlife : Gaps in 3400

knowledge. N. Z. J. Ecol. 30, 157–167. 3401

Hobbs, R.J., Saunders, D.A., 1991. Re-integrating fragmented landscapes-a preliminary 3402

framework for the Western Australian wheatbelt. J. Environ. Manage. 33, 161–167. 3403

https://doi.org/10.1016/S0301-4797(05)80092-1 3404

Hogan, F., Burridge, C., Cooke, R., Norman, J., 2007. Isolation and characterization of 3405

microsatellite loci to DNA fingerprint the Powerful Owl (Ninox strenua). Mol. Ecol. 3406

Notes 7, 1305–1307. https://doi.org/10.1111/j.1471-8286.2007.01864.x 3407

Hogan, F.E., Cooke, R., 2010. Insights into the breeding behaviour and dispersal of the 3408

Powerful Owl (Ninox strenua) through the collection of shed feathers. Emu 110, 178–3409

184. https://doi.org/10.1071/MU09116 3410

Hogan, F.E., Cooke, R., Norman, J.A., 2009. Reverse ascertainment bias in microsatellite 3411

allelic diversity in owls (Aves, Strigiformes). Conserv. Genet. 10, 635–638. 3412

https://doi.org/10.1007/s10592-008-9595-0 3413

Hogg, M.J., Skegg, P.D.G., 1961. Moreporks in a Nesting Box. Notornis 9, 133–134. 3414

Hollings, T., Jones, M., Mooney, N., McCallum, H., 2013. Wildlife disease ecology in changing 3415

landscapes: Mesopredator release and toxoplasmosis. Int. J. Parasitol. Parasites Wildl. 3416

2, 110–118. https://doi.org/10.1016/j.ijppaw.2013.02.002 3417

Horak, K., Fisher, P.M., Hopkins, B.M., 2018. Pharmacokinetics of Anticoagulant 3418

Rodenticides in Target and Non-target Organisms, in: Emerging Topics in Ecotoxicology 3419

161

5. pp. 87–108. https://doi.org/10.1007/978-3-319-64377-9 3420

Houlahan, J.E., Findlay, C.S., Schmidt, B.R., Meyer, A.H., Kuzmin, S.L., 2000. Quantitative 3421

evidence for global amphibian population declines. Nature 404, 752–755. 3422

https://doi.org/10.1038/35008052 3423

How, R.A., Dell, J., 1994. The zoogeographic significance of urban bushland remnants to 3424

reptiles in the Perth region, Western Australia. Pacific Conserv. Biol. 1, 132–140. 3425

https://doi.org/10.1071/pc940132 3426

Hubisz, M.J., Falush, D., Stephens, M., Pritchard, J.K., 2009. Inferring weak population 3427

structure with the assistance of sample group information. Mol. Ecol. Resour. 9, 1322–3428

1332. https://doi.org/10.1111/j.1755-0998.2009.02591.x 3429

Hughes, J., Sharp, E., Taylor, M.J., Melton, L., Hartley, G., 2013. Monitoring agricultural 3430

rodenticide use and secondary exposure of raptors in Scotland. Ecotoxicology 22, 974–3431

984. https://doi.org/10.1007/s10646-013-1074-9 3432

Hussey, P., 1997. Nest boxes for native wildlife. Wildl. Notes 1–6. 3433

Isaac, B., Cooke, R., Ierodiaconou, D., White, J., 2014a. Does urbanization have the potential 3434

to create an ecological trap for powerful owls (Ninox strenua)? Biol. Conserv. 176, 1–3435

11. https://doi.org/10.1016/j.biocon.2014.04.013 3436

Isaac, B., Cooke, R., Simmons, D., Hogan, F., 2008. Predictive mapping of powerful owl 3437

(Ninox strenua) breeding sites using Geographical Information Systems (GIS) in urban 3438

Melbourne, Australia. Landsc. Urban Plan. 84, 212–218. 3439

https://doi.org/10.1016/j.landurbplan.2007.08.002 3440

Isaac, B., White, J., Ierodiaconou, D., Cooke, R., 2014b. Urban to forest gradients : Suitability 3441

for hollow bearing trees and implications for obligate hollow nesters. Austral Ecol. 3442

doi:10.111, 1–10. https://doi.org/10.1111/aec.12164 3443

Isaac, B., White, J., Ierodiaconou, D., Cooke, R., 2013. Response of a cryptic apex predator to 3444

a complete urban to forest gradient. Wildl. Res. 40, 427–436. 3445

https://doi.org/10.1071/WR13087 3446

James, D., 1997. Rat-baiting in cane fields & poisoning of wildlife. Boobook 18, 5. 3447

162

Joermann, G., 1998. A review of secondary-poisoning studies with rodenticides. Bull. 3448

OEPP/EPPO Bull. 28, 157–176. https://doi.org/10.1111/j.1365-2338.1998.tb00717.x 3449

Johnson, D.H., 1980. The comparison of usage and availability measurements for evaluating 3450

resource preference. Ecology 61, 65–71. https://doi.org/10.2307/1937156 3451

Johnstone, B.R.E., Kirkby, T., Mannion, M., 2015. Trials on the use and effectiveness of 3452

artificial nest hollows for Carnaby’s Cockatoo at Cataby, Western Australia. West. Aust. 3453

Nat. 29, 250–262. 3454

Johnstone, R.E., Kirkby, T., 2007. Feral European honey bees: a major threat to cockatoos 3455

and other tree hollow users. West. Aust. Nat. 25, 252-254. 3456

Jolly, S.E., Eason, C.T., Frampton, C., Gumbrell, R.C., 1994. The anticoagulant pindone causes 3457

liver damage in the brushtail possum (Trichosurus vulpecula). Aust. Vet. J. 71, 220. 3458

Jules, E.S., Shahani, P., 2003. A broader ecological context to habitat fragmentation: Why 3459

matrix habitat is more important than we thought. J. Veg. Sci. 14, 459–464. 3460

https://doi.org/10.1111/j.1654-1103.2003.tb02172.x 3461

Kan, I., Motro, Y., Horvitz, N., Kimhi, A., Leshem, Y., Yom-Tov, Y., Nathan, R., 2013. 3462

Agricultural Rodent Control Using Barn Owls: Is It Profitable? Am. J. Agric. Econ. 96, 3463

733–752. https://doi.org/10.1093/ajae/aat097 3464

Kaukeinen, D.E., Spragins, C.W., Hobson, J.F., 2000. Risk-benefit considerations in evaluating 3465

commensal anticoagulant rodenticide impacts to wildlife. Proc. Vertebr. Pest Conf. 19, 3466

245–256. 3467

Kavanagh, R., Bamkin, K., 1995. Distribution of nocturnal forest birds and mammals in 3468

relation to the logging mosaic in south-eastern New South Wales, Australia. Biol. 3469

Conserv. 71, 41–53. 3470

Kavanagh, R., Debus, S., Tweedie, T., Webster, R., 1995. Distribution of Nocturnal Forest 3471

Birds and Mammals in North-Eastern New South Wales: Relationships With 3472

Environmental Variables and Management History. Wildl. Res. 22, 359–377. 3473

https://doi.org/10.1071/WR9950359 3474

Kavanagh, R., Peake, P., 1993. Distribution and habitats of nocturnal forest birds in south-3475

163

eastern New South Wales, Australian Raptor Studies. Australasian Raptor Association, 3476

Royal Australasian Ornithologists Union, Melbourne. 3477

Kavanagh, R.P., Murray, M., 1996. Home Range, Habitat and Behaviour of the Masked Owl 3478

Tyto novaehollandiae near Newcastle, New South Wales. Emu 96, 250–257. 3479

Kavanagh, R.P., Stanton, M.A., 2002. Response to habitat fragmentation by the Powerful 3480

Owl (Ninox strenua), Sooty Owl (Tyto tenebricosa) and Masked Owl (Tyto 3481

novaehollandiae) and other nocturnal fauna in south-eastern Australia, in: Newton, I., 3482

Kavanagh, R., Olsen, J., Taylor, I. (Eds.), Ecology and Conservation of Owls. CSIRO 3483

Publishing, Collingwood, Victoria, pp. 265–277. 3484

Kay, B.J., Twigg, L.E., Nicol, H.I., 1994. The Strategic Use of Rodenticides against House Mice 3485

(Mus domesticus) prior to Crop Invasion. Wildl. Res. 11–19. 3486

Khalifa, B.A.A., Salem, F.M.S., Menha, M.A., 1992. Embryotoxic and Teratogenic Effects of 3487

Flocoumafen in Chick Embryos and White Rats. J. Appl. Anim. Res. 2, 81–85. 3488

https://doi.org/10.1080/09712119.1992.9705936 3489

King, K.C., McLaughlin, J.D., Gendron, A.D., Pauli, B.D., Giroux, I., Rondeau, B., Boily, M., 3490

Juneau, P., Marcogliese, D.J., 2007. Impacts of agriculture on the parasite communities 3491

of northern leopard frogs (Rana pipiens) in southern Quebec, Canada. Parasitology 134, 3492

2063–2080. https://doi.org/10.1017/S0031182007003277 3493

Kirkpatrick, C.E., Colvin, B.A., Dubey, J.P., 1990. Toxoplasma gondii antibodies in common 3494

barn-owls (Tyto alba) and pigeons (Columba livia) in New Jersey. Vet. Parasitol. 36, 3495

177–180. https://doi.org/10.1016/0304-4017(90)90108-N 3496

Klein, Á., Nagy, T., Csörgő, T., Mátics, R., 2007. Exterior nest-boxes may negatively affect 3497

Barn Owl Tyto alba survival: an ecological trap. Bird Conserv. Int. 3498

https://doi.org/10.1017/S0959270907000792 3499

Klemens, M.W., Thorbjarnarson, J.B., 1995. Reptiles as a food resource. Biodivers. Conserv. 3500

4, 281–298. 3501

Knopper, L.D., Mineau, P., Walker, L.A., Shore, R.F., 2007. Bone density and breaking 3502

strength in UK raptors exposed to second generation anticoagulant rodenticides. Bull. 3503

164

Environ. Contam. Toxicol. 78, 249–51. https://doi.org/10.1007/s00128-007-9122-x 3504

König, C., Weick, F., Becking, J.H., 1999. Owls. A Guide to the Owls of the World. Pica Press, 3505

Sussex. 3506

Kopanke, J.H., Horak, K.E., Musselman, E., Miller, C.A., Bennett, K., Olver, C.S., Volker, S.F., 3507

Vandewoude, S., Bevins, S.N., 2018. Effects of Low-level Brodifacoum Exposure on the 3508

Feline Immune Response. Sci. Rep. 8, 1–13. https://doi.org/10.1038/s41598-018-3509

26558-3 3510

Kopelman, N.M., Mayzel, J., Jakobsson, M., Rosenberg, N.A., Mayrose, I., 2015. CLUMPAK: a 3511

program for identifying clustering modes and packaging population structure 3512

inferences across K. Mol. Ecol. Resour. 15, 1179–1191. https://doi.org/doi: 3513

10.1111/1755-0998.12387 3514

Krawiec, J., Krauss, S.L., Davis, R.A., Spencer, P.B.S., 2015. Weak genetic structuring suggests 3515

historically high genetic connectivity among recently fragmented urban populations of 3516

the scincid lizard, Ctenotus fallens. Aust. J. Zool. 63, 279. 3517

https://doi.org/10.1071/zo15022 3518

Kusbeci, O.Y., Miman, O., Yaman, M., Aktepe, O.C., Yazar, S., 2011. Could Toxoplasma gondii 3519

have any role in Alzheimer disease? Alzheimer Dis. Assoc. Disord. 25, 1–3. 3520

Laakso, S., Suomalainen, K., Koivisto, S., 2010. Literature review on residues of anticoagulant 3521

rodenticides in nontarget animals. Copenhagen. 3522

Lafferty, K.D., 2006. Can the common brain parasite, Toxoplasma gondii, influence human 3523

culture? Proc. R. Soc. London B Biol. Sci. 273, 2749–2755. 3524

https://doi.org/10.1098/rspb.2006.3641 3525

LaHaye, W.S., Gutiérrez, R.J., Dunk, J.R., 2001. Natal dispersal of the spotted owl in southern 3526

California: dispersal profile of an insular population. Condor 103, 691–700. 3527

https://doi.org/10.2307/1370102 3528

Lambrechts, M.M., Wiebe, K.L., Sunde, P., Solonen, T., Sergio, F., Roulin, A., Møller, A.P., 3529

López, B.C., Fargallo, J. a., Exo, K.M., Dell’Omo, G., Costantini, D., Charter, M., Butler, 3530

M.W., Bortolotti, G.R., Arlettaz, R., Korpimäki, E., 2012. Nest box design for the study of 3531

165

diurnal raptors and owls is still an overlooked point in ecological, evolutionary and 3532

conservation studies: A review. J. Ornithol. 153, 23–34. 3533

https://doi.org/10.1007/s10336-011-0720-3 3534

Langford, K.H., Reid, M., Thomas, K. V, 2013. The occurrence of second generation 3535

anticoagulant rodenticides in non-target raptor species in Norway. Sci. Total Environ. 3536

450–451, 205–208. https://doi.org/10.1016/j.scitotenv.2013.01.100 3537

Leake, B., 1962. Eastern Wheatbelt Wildlife: Experiences of a WA Naturalist. Docket Book 3538

Company, Perth. 3539

Leck, C.F., 1979. Avian Extinctions in an Isolated Tropical Wet-Forest Preserve, Ecuador. Auk 3540

96, 343–352. 3541

Lemus, J.A., Bravo, C., García-Montijano, M., Palacín, C., Ponce, C., Magaña, M., Alonso, J.C., 3542

2011. Side effects of rodent control on non-target species: Rodenticides increase 3543

parasite and pathogen burden in great bustards. Sci. Total Environ. 409, 4729–4734. 3544

https://doi.org/10.1016/j.scitotenv.2011.07.007 3545

León-Ortega, M., GonzáLez-Wangoemert, M., Martfinez, J.E., Calvo, J.F., 2014. Spatial 3546

patterns of genetic diversity in mediterranean eagle owl Bubo bubo populations. 3547

Ardeola 61, 45–62. https://doi.org/10.13157/arla.61.1.2014.45 3548

Levins, R., 1969. Some demographic and genetic consequences of environmental 3549

heterogeneity for biological control. Bull. Entomol. Soc. Am. 15, 237–240. 3550

Liddelow, G.L., Wheeler, I.B., Kavanagh, R.P., 2002. Owls in the South-West Forests of 3551

Western Australia, in: Newton, I., Kavanagh, R., Olsen, J., Taylor, I. (Eds.), Ecology and 3552

Conservation of Owls. CSIRO Publishing, Collingwood, Victoria, pp. 233-241. 3553

Lindenmayer, D.B., Cunningham, R.B., Donnelly, C.F., 1997. Decay and collapse of trees with 3554

hollows in eastern Australian forests: Impacts on arboreal marsupials. Ecol. Appl. 7, 3555

625–641. https://doi.org/10.1890/1051-0761(1997)007[0625:DACOTW]2.0.CO;2 3556

Lindsay, D.S., Smith, P.C., Hoerr, F.J., Blagburn, B.L., 1993. Prevalence of encysted 3557

Toxoplasma gondii in raptors from Alabama. J. Parasitol. 79, 870–873. 3558

https://doi.org/10.2307/3283724 3559

166

Littin, K.E., O’Connor, C.E., Gregory, N.G., Mellor, D.J., Eason, C.T., 2002. Behaviour, 3560

coagulopathy and pathology of brushtail possums (Trichosurus vulpecula) poisoned 3561

with brodifacoum. Wildl. Res. 29, 259–267. 3562

Liu, Q., Wang, Z.-D., Huang, S.-Y., Zhu, X.-Q., 2015. Diagnosis of toxoplasmosis and typing of 3563

Toxoplasma gondii. Parasites and Vectors 8, 1–14. https://doi.org/10.1186/s13071-3564

015-0902-6 3565

Lohr, M.T., 2018. Anticoagulant rodenticide exposure in an Australian predatory bird 3566

increases with proximity to developed habitat. Sci. Total Environ. 643, 134–144. 3567

https://doi.org/10.1016/j.scitotenv.2018.06.207 3568

Lohr, M.T., Davis, R.A., 2018. Anticoagulant rodenticide use, non-target impacts and 3569

regulation: A case study from Australia. Sci. Total Environ. 634, 1372–1384. 3570

https://doi.org/10.1016/j.scitotenv.2018.04.069 3571

Loman, J., 2006. Does nest site availability limit the density of hole nesting birds in small 3572

woodland patches? Web Ecol. 6, 37–43. 3573

Lopes, A.P., Sargo, R., Rodrigues, M., Cardoso, L., 2011. High seroprevalence of antibodies to 3574

Toxoplasma gondii in wild animals from Portugal. Parasitol. Res. 108, 1163–1169. 3575

https://doi.org/10.1007/s00436-010-2158-6 3576

López-Perea, J.J., Camarero, P.R., Molina-lópez, R.A., Parpal, L., Obón, E., Solá, J., Mateo, R., 3577

2015. Interspecific and geographical differences in anticoagulant rodenticide residues 3578

of predatory wildlife from the Mediterranean region of Spain. Sci. Total Environ. 511, 3579

259–67. https://doi.org/10.1016/j.scitotenv.2014.12.042 3580

Love, D., Kwok, O.C., Verma, S.K., Dubey, J.P., Bellah, J., 2016. Antibody Prevalence and 3581

Isolation of Viable Toxoplasma gondii from Raptors in the Southeastern USA. J. Wildl. 3582

Dis. 52, 653–656. https://doi.org/10.7589/2015-10-269 3583

MacArthur, R.H., Wilson, E.O., MacArthur, W., 1967. The theory of island biogeography, 3584

Princeton University Press Princeton USA Major RE Kendal CE. 3585

https://doi.org/10.2307/1796430 3586

Mack, D.B., 1965. Nest predation of Boobook Owl. South Aust. Ornithol. 24, 51. 3587

167

Markovitz, A.A., Simanek, A.M., Yolken, R.H., Galea, S., Koenen, K.C., Chen, S., Aiello, A.E., 3588

2015. Toxoplasma gondii and anxiety disorders in a community-based sample. Brain. 3589

Behav. Immun. 43, 192–197. https://doi.org/10.1016/j.bbi.2014.08.001 3590

Martin, G.R., Kirkpatrick, W., King, D., Robertson, I., Hood, P., Sutherland, J., 1994. 3591

Assessment of the Potential Toxicity of an Anticoagulant, Pindone (2-Pivalyl-1,3-3592

Indandione), to Some Australian Birds. Wildl. Res. 21, 85–93. 3593

https://doi.org/10.1071/WR9940085 3594

Martin, K., Eadie, J.M., 1999. Nest webs: A community-wide approach to the management 3595

and conservation of cavity-nesting forest birds. For. Ecol. Manage. 115, 243–257. 3596

https://doi.org/10.1016/S0378-1127(98)00403-4 3597

Martínez-Padilla, J., López-Idiáquez, D., López-Perea, J.J., Mateo, R., Paz, A., Viñuela, J., 3598

2016. A negative association between bromadiolone exposure and nestling body 3599

condition in common kestrels: management implications for vole outbreaks. Pest 3600

Manag. Sci. 73, 364–370. https://doi.org/10.1002/ps.4435 3601

Mason, R.W., Hartley, W.J., Dubey, J.P., 1991. Lethal toxoplasmosis in a little penguin 3602

(Eudyptula minor) from Tasmania. J. Parasitol. 77, 328. 3603

https://doi.org/10.2307/3283109 3604

Masters, J.R., Milhinch, A.L., 1974. Birds of the Shire of Northam, about 100 km east of 3605

Perth. Emu. https://doi.org/10.1071/MU974228 3606

Masuda, B.M., Fisher, P., Jamieson, I.G., 2014. Anticoagulant rodenticide brodifacoum 3607

detected in dead nestlings of an insectivorous passerine. N. Z. J. Ecol. 38, 110–115. 3608

Matlack, G.R., 1993. Microenvironment variation within and among forest edge sites in the 3609

eastern United States. Biol. Conserv. 66, 185–194. https://doi.org/10.1016/0006-3610

3207(93)90004-K 3611

Mcdonald, P.J., Pavey, C.R., 2014. Exploiting boom times . Southern Boobook Owl Ninox 3612

novaseelandiae diet during a rodent irruption in central Australia. Aust. Zool. 37, 234–3613

237. 3614

Mcdonald, R.A., Harris, S., Turnbull, G., Brown, P., Fletcher, M., 1998. Anticoagulant 3615

168

rodenticides in stoats (Mustela erminea) and weasels (Mustela nivalis) in England. 3616

Environ. Pollut. 103, 17–23. 3617

McFadden, D., 1978. Quantitative Methods for Analysing Travel Behaviour of Individuals: 3618

Some Recent Developments, in: Hensher, D.A., Stopher, P.R. (Eds.), Behavioural Travel 3619

Modelling. Croom Helm, London, p. 307. 3620

McIlroy, J.C., 2014. New techniques for an old problem-recent advances in feral pig control 3621

in Australia. J. Mt. Ecol. 3, 241–244. 3622

McIlroy, J.C., Braysher, M., Saunders, G.R., 1989. Effectiveness of a Warfarin-poisoning 3623

Campaign Against Feral Pigs, Sus scrofa, in Namadgi National Park, A.C.T. Aust. Wildl. 3624

Res. 16, 195–202. https://doi.org/10.1071/WR9890195 3625

Mcmillin, S.C., Hosea, R.C., Finlayson, B.F., Cypher, B.L., Mekebri, A., 2008. Anticoagulant 3626

Rodenticide Exposure in an Urban Population of the San Joaquin Kit Fox, in: Timm, 3627

R.M., Madon, M.B. (Eds.), Proceedings of the 23rd Vertebrate Pest Conference. pp. 3628

163–165. 3629

McNabb, E., Keating, M., 2008. Helping Powerful Owls. Wingspan 18, 30–33. 3630

McNabb, E.D., Greenwood, J., 2011. A powerful Owl disperses into town and uses an 3631

artificial nest-box. Aust. F. Ornithol. 28, 65–75. 3632

Mecca, J.N., Meireles, L.R., de Andrade, H.F., 2011. Quality control of Toxoplasma gondii in 3633

meat packages: Standardization of an ELISA test and its use for detection in rabbit meat 3634

cuts. Meat Sci. 88, 584–589. https://doi.org/10.1016/j.meatsci.2011.01.016 3635

Meek, P.D., Hawksby, R.J., Ardler, A., Hudson, M., Tuckey, K.D., 2011. Eradication of Black 3636

Rats Rattus rattus L . from Bowen Island, Jervis Bay NSW. Aust. Zool. 35, 560–568. 3637

Meerburg, B.G., Kijlstra, A., 2009. Changing climate-changing pathogens: Toxoplasma gondii 3638

in North-Western Europe. Parasitol. Res. 105, 17–24. https://doi.org/10.1007/s00436-3639

009-1447-4 3640

Mendenhall, V.M., Pank, L.F., 1980. Secondary Poisoning of Owls by Anticoagulant 3641

Rodenticides. Wildl. Soc. Bull. 8, 311–315. 3642

Merton, D.V., 1987. Eradication of rabbits from Round Island, Mauritius: a conservation 3643

169

success story. Dodo 24, 19–24. 3644

Mikaelian, I., Dubey, J.P., Martineau, D., 1997. Severe hepatitis resulting from toxoplasmosis 3645

in a barred owl (Strix varia) from Québec, Canada. Avian Dis. 41, 738–740. 3646

Milledge, D., Palmer, C., Nelson, J., 1991. Barometers of change: the distribution of large 3647

owls and gliders in Mountain Ash forests of the Victorian Central Highlands and their 3648

potential as management indicators, in: Lunney, D. (Ed.), Conservation of Australia’s 3649

Forest Fauna. Royal Zoological Society of New South Wales, Mosman, pp. 53–65. 3650

Miller, N.L., Frenkel, J.K., Dubey, J.P., 1972. Oral infections with Toxoplasma cysts and 3651

oocysts in felines, other mammals, and in birds. J. Parasitol. 58, 928–37. 3652

Miller, S.A., Dykes, D.D., Polesky, H.F., 1988. A simple salting out procedure for extracting 3653

DNA from human nucleated calls. Nucleic Acids Res. 16, 1215. 3654

https://doi.org/10.1016/B978-0-7506-7152-1.50011-4 3655

Miman, O., Kusbeci, O.Y., Aktepe, O.C., Cetinkaya, Z., 2010a. The probable relation between 3656

Toxoplasma gondii and Parkinson’s disease. Neurosci. Lett. 475, 129–131. 3657

https://doi.org/10.1016/j.neulet.2010.03.057 3658

Miman, O., Mutlu, E.A., Ozcan, O., Atambay, M., Karlidag, R., Unal, S., 2010b. Is there any 3659

role of Toxoplasma gondii in the etiology of obsessive-compulsive disorder? Psychiatry 3660

Res. 177, 263–265. https://doi.org/10.1016/j.psychres.2009.12.013 3661

Monreal, M., Olive, A., Lafoz, E., del Rio, L., 1991. Heparins, coumarin, and bone density. 3662

Lancet 338, 706. 3663

Mooney, N., 2017. Risks of Anticoagulant Rodenticides to Tasmanian Raptors. Tasmanian 3664

Bird Rep. 38, 17–25. 3665

Mooney, P.A., Pedler, L.P., 2005. Recovery Plan for the South Australian subspecies of the 3666

Glossy Black-Cockatoo (Calyptorhynchus lathami halmaturinus): 2005-2010. Adelaide. 3667

Morgan, D.K.J., Styche, A., 2012. Results of a community-based acoustic survey of ruru 3668

(moreporks) in Hamilton city. Notornis 59, 123–129. 3669

Moro, D., 2001. Evaluation and cost-benefits of controlling house mice (Mus domesticus) on 3670

islands: an example from Thevenard Island, Western Australia. Biol. Conserv. 99, 355–3671

170

364. 3672

Morris, K.D., 2002. The eradication of the black rat (Rattus rattus) on Barrow and adjacent 3673

islands off the north-west coast of Western Australia, in: Veitch, C.R., Clout, M.N. 3674

(Eds.), Turning the Tide : The Eradication of Invasive Species. IUCN Species Survival 3675

Commission, Gland, Switzerland, pp. 219–225. 3676

Morton, E., Forman, L., Braun, M., 1990. Extrapair fertilizations and the evolution of colonial 3677

breeding in purple martins. Auk 107, 275–283. 3678

Mosterd, J.J., Thijssen, H.H.W., 1991. The long-term effects of the rodenticide, brodifacoum, 3679

on blood coagulation and vitamin K metabolism in rats. Br. J. Pharmacol. 104, 531–535. 3680

Moura, R.F., Dawson, D.A., Nogueira, D.M., 2017. The use of microsatellite markers in 3681

neotropical studies of wild birds: A literature review. An. Acad. Bras. Cienc. 89, 145–3682

154. https://doi.org/10.1590/0001-3765201620160378 3683

Murray, M., 2017. Anticoagulant rodenticide exposure and toxicosis in four species of birds 3684

of prey in Massachusetts, USA, 2012 – 2016, in relation to use of rodenticides by pest 3685

management professionals. Ecotoxicology 26, 1041–1050. 3686

https://doi.org/10.1007/s10646-017-1832-1 3687

Naim, M., Hafidzi, M.N., Sudharto, P.S., Caliman, J.P., 2012. Detection of rodenticide 3688

residues for raptor conservation, in: 3rd International Conference on Oil Palm and 3689

Environment (ICOPE). Bali, Indonesia, pp. 22–24. 3690

Naim, M., Noor, H.M., Kasim, A., Abu, J., 2010. Growth performance of nestling barn owls, 3691

Tyto alba javanica in rat baiting area in Malaysia. ARPN J. Agric. Biol. Sci. 5, 1–13. 3692

National Registration Authority For Agricultural and Veterinary Chemicals, 2002. The NRA 3693

Review of PINDONE. Canberra, Australia. 3694

New South Wales Department of Primary Industries, 2011. Mouse monitoring and baiting 3695

[WWW Document]. URL 3696

http://www.dpi.nsw.gov.au/__data/assets/pdf_file/0008/387719/mouse-monitoring-3697

and-baiting.pdf (accessed 8.21.17). 3698

New South Wales Government: Department of Primary Industries, 2017. Mouse baiting 3699

171

FAQs [WWW Document]. URL http://www.dpi.nsw.gov.au/biosecurity/vertebrate-3700

pests/pest-animals-in-nsw/mouse-and-mice-plagues/mouse-baiting-faqs#Rodenticides 3701

(accessed 8.24.17). 3702

Newton, I., 2004. The recent declines of farmland bird populations in Britain: An appraisal of 3703

causal factors and conservation actions. Ibis (Lond. 1859). 146, 579–600. 3704

https://doi.org/10.1111/j.1474-919X.2004.00375.x 3705

Newton, I., 1994. The role of nest sites in limiting the numbers of hole-nesting birds: A 3706

review. Biol. Conserv. 70, 265–276. https://doi.org/10.1016/0006-3207(94)90172-4 3707

Newton, I., Dale, L., Finnie, J., Freestone, P., Wright, J., Wyatt, C., Wyllie, I., 1998. Wildlife 3708

and pollution: 1997/98 Annual Report. 3709

Newton, I., Shore, R.F., Wyllie, I., Birks, J.D.S., Dale., L., 1999. Empirical evidence of side-3710

effects of rodenticides on some predatory birds and mammals, in: Cowan, D.P., Feare, 3711

C.J. (Eds.), Advances in Vertebrate Pest Management. Filander Verlag, Fürth, pp. 347–3712

367. 3713

Newton, I., Wyllie, I., Freestone, P., 1990. Rodenticides in British barn owls. Environ. Pollut. 3714

68, 101–117. https://doi.org/10.1016/0269-7491(90)90015-5 3715

Ngoungou, E.B., Bhalla, D., Nzoghe, A., Dardé, M.-L., Preux, P.-M., 2015. Toxoplasmosis and 3716

Epilepsy — Systematic Review and Meta Analysis. PLoS Negl. Trop. Dis. 9, e0003525. 3717

https://doi.org/10.1371/journal.pntd.0003525 3718

Ninox boobook [WWW Document], 2018. . IUCN Red List Threat. Species. URL 3719

http://www.iucnredlist.org/details/62023787/0 (accessed 6.7.18). 3720

Nogeire, T.M., Lawler, J.J., Schumaker, N.H., Cypher, B.L., Phillips, S.E., 2015. Land use as a 3721

driver of patterns of rodenticide exposure in modeled kit fox populations. PLoS One 10, 3722

E0133351. https://doi.org/10.1371/journal.pone.0133351 3723

Obendorf, D., Statham, P., Driessen, M., 1996. Detection of agglutinating antibodies to 3724

Toxoplasma gondii in sera from free-ranging eastern barred bandicoots (Perameles 3725

gunnii). J. Wildl. Dis. 32, 623–626. 3726

Oliver, A.J., Wheeler, S.H., Gooding, C.D., 1982. Field Evaluation of 1080 and Pindone Oat 3727

172

Bait, and the Possible Decline in Effectiveness of Poison Baiting for the Control of the 3728

Rabbit, Oryctolagus cuniculus. Aust. Wildl. Res. 9, 125–134. 3729

Olsen, Debus, S.J.S., 2013. Forum- Do Tasmanian southern boobooks migrate? Aust. F. 3730

Ornithol. 30, 106–108. 3731

Olsen, J., 2011a. What is a Southern Boobook?, in: Australian High Country Owls. CSIRO 3732

PUBLISHING, Collingwood, Victoria, pp. 15–19. 3733

Olsen, J., 2011b. Southern Boobook calls, in: Australian High Country Owls. CSIRO 3734

Publishing, Collingwood, Victoria, pp. 119–127. 3735

Olsen, J., Debus, S.J.S., Judge, D., 2013. Declining little eagles Hieraaetus morphnoides and 3736

increasing rabbit numbers near canberra: Is secondary poisoning by pindone the 3737

problem? Corella 37, 33–35. 3738

Olsen, J., Downs, J. a., Tucker, T., Trost, S., 2011. Home-Range Size and Territorial Calling of 3739

Southern Boobooks (Ninox novaeseelandiae) in Adjacent Territories. J. Raptor Res. 45, 3740

136–142. https://doi.org/10.3356/JRR-10-92.1 3741

Olsen, J., Fuentes, E., Rose, A.B., Trost, S., 2006. Food and Hunting of Eight Breeding Raptors 3742

Near Canberra, 1990-1994. Aust. F. Ornithol. 23, 77–95. 3743

Olsen, J., Judge, D., Trost, S., 2015. Ageing southern boobook nestlings and fledglings. 3744

Corella 39, 77. 3745

Olsen, J., Taylor, S., 2001. Winter home range of an adult female southern boobook, Ninox 3746

novaeseelandieae in suburban Canberra. Aust. Bird Watch. 19, 109–114. 3747

Olsen, J., Trost, S., 2015. Are Southern Boobooks Declining? Canberra Bird Notes 40, 270–3748

271. 3749

Olsen, J., Trost, S., 2007. Duelling and nest failures in Southern Boobooks Ninox 3750

novaeseelandiae. Aust. F. Ornithol. 24, 13–25. 3751

Olsen, P., Debus, S., Czechura, G., Mooney, N., 1990. Comparative feeding ecology of the 3752

Grey Goshawk Accipiter novaehollandiae and Brown Goshawk Accipiter fasciatus. Aust. 3753

Bird Watch. 13, 178–192. 3754

173

Olsen, P., Rae, S., 2017. Invalid Evidence for Purported ‘Collapse’ in the Number of Breeding 3755

Little Eagles in the Australian Capital Territory. Canberra Bird Notes 42, 245–249. 3756

Olsen, P.D., 1996. Re-establishment of an endangered subspecies: the Norfolk Island 3757

Boobook Owl Ninox novaeseelandiae undulata. Bird Conserv. Int. 6, 63–80. 3758

https://doi.org/10.1017/S0959270900001313 3759

Olsson, M., Wapstra, E., Swan, G., Snaith, E., Clarke, R., Madsen, T., 2005. Effects of long-3760

term fox baiting on species composition and abundance in an Australian lizard 3761

community. Austral Ecol. 30, 907–913. https://doi.org/10.1111/j.1442-3762

9993.2005.01534.x 3763

Omote, K., Nishida, C., Takenaka, T., Saito, K., Shimura, R., Fujimoto, S., Sato, T., Masuda, R., 3764

2015. Recent fragmentation of the endangered Blakiston’s fish owl (Bubo blakistoni) 3765

population on Hokkaido Island, Northern Japan, revealed by mitochondrial DNA and 3766

microsatellite analyses. Zool. Lett. 1, 16. https://doi.org/10.1186/s40851-015-0014-3 3767

Osborne, N.J., Cairns, R., Dawson, A.H., Chitty, K.M., Buckley, N.A., 2017. Epidemiology of 3768

coronial deaths from pesticide ingestion in Australia. Int. J. Hyg. Environ. Health 220, 3769

478–484. https://doi.org/10.1016/j.ijheh.2017.01.009 3770

Owen, J.C., 2011. Collecting, processing, and storing avian blood: A review. J. F. Ornithol. 82, 3771

339–354. https://doi.org/10.1111/j.1557-9263.2011.00338.x 3772

Ozanne-Smith, J., Day, L., Parsons, B., Tibballs, J., Dobbin, M., 2001. Childhood poisoning: 3773

Access and prevention. J. Paediatr. Child Health 37, 262–265. 3774

Pacioni, C., Eden, P., Reiss, A., Ellis, T., Knowles, G., Wayne, A.F., 2015. Disease hazard 3775

identification and assessment associated with wildlife population declines. Ecol. 3776

Manag. Restor. 16, 142–152. https://doi.org/10.1111/emr.12155 3777

Paltridge, R., 2002. The diets of cats, foxes and dingoes in relation to prey availability the the 3778

Tanami Desert, Northern Territory. Wildl. Res. 29, 389–403. 3779

Pan, S., Thompson, R.C.A., Grigg, M.E., Sundar, N., Smith, A., Lymbery, A.J., 2012. Western 3780

Australian marsupials are multiply infected with genetically diverse strains of 3781

Toxoplasma gondii. PLoS One 7, e45147. 3782

174

https://doi.org/10.1371/journal.pone.0045147 3783

Parameswaran, N., 2008. Toxoplasma gondii in Australian marsupials. Murdoch University. 3784

Parameswaran, N., O’Handley, R.M., Grigg, M.E., Wayne, A., Thompson, R.C.A., 2009. 3785

Vertical transmission of Toxoplasma gondii in Australian marsupials. Parasitology 136, 3786

939–944. https://doi.org/10.1017/S0031182009006453 3787

Park, B., Scott, A., Wilson, A., Haynes, B., Breckenridge, A., 1984. Plasma disposition of 3788

vitamin K1 in relation to anticoagulant poisoning. Br. J. Clin. Pharmacol. 18, 655–662. 3789

https://doi.org/10.1111/j.1365-2125.1984.tb02526.x 3790

Parshad, V.R., 1999. Rodent Control in India. Integr. Pest Manag. Rev. 4, 97–126. 3791

https://doi.org/10.1023/A:1009622109901 3792

Parsons, B.J., Day, L.M., Ozanne-Smith, J., Dobbin, M., 1996. Rodenticide poisoning among 3793

children. Aust. New Zeal. J. Public Heal. 20, 488–492. 3794

Patton, S., Johnson, S.L., Loeffler, D.G., Wright, B.G., Jensen, J.M., 1986. Epizootic of 3795

toxoplasmosis in kangaroos, wallabies, and potaroos: possible transmission via 3796

domestic cats. J. Am. Vet. Med. Assoc. 189, 1166–1169. 3797

Peakall, R., Smouse, P.E., 2012. GenALEx 6.5: Genetic analysis in Excel. Population genetic 3798

software for teaching and research-an update. Bioinformatics 28, 2537–2539. 3799

https://doi.org/10.1093/bioinformatics/bts460 3800

Peakall, R., Smouse, P.E., 2006. GENALEX 6: genetic analysis in Excel. Population genetic 3801

software for teaching and research. Mol. Ecol. Notes 6, 288–295. 3802

https://doi.org/10.1111/j.1471-8286.2005.01155.x 3803

Petty, S.J., 1992. Ecology of the tawny owl Strix aluco in the spruce forests of 3804

Northumberland and Argyll. Open University. 3805

Pither, J., Taylor, P.D., 1998. An experimental assessment of landscape connectivity. Oikos 3806

83, 166–174. https://doi.org/10.2307/3546558 3807

Pitt, W.C., Berentsen, A.R., Shiels, A.B., Volker, S.F., Eisemann, J.D., Wegmann, A.S., Howald, 3808

G.R., 2015. Non-target species mortality and the measurement of brodifacoum 3809

rodenticide residues after a rat (Rattus rattus) eradication on Palmyra Atoll, tropical 3810

175

Pacific. Biol. Conserv. 185, 36–46. https://doi.org/10.1016/j.biocon.2015.01.008 3811

Pitt, W.C., Higashi, M., Primus, T.M., 2011. The effect of cooking on diphacinone residues 3812

related to human consumption of feral pig tissues of feral pig tissues. Food Chem. 3813

Toxicol. 49, 2030–2034. https://doi.org/10.1016/j.fct.2011.05.014 3814

Poché, R.M., 1988. Rodent tissue residue and secondary hazard studies with bromadiolone. 3815

EPPO Bull. 18, 323–330. https://doi.org/10.1111/j.1365-2338.1988.tb00382.x 3816

Prandota, J., 2010. Autism spectrum disorders may be due to cerebral toxoplasmosis 3817

associated with chronic neuroinflammation causing persistent hypercytokinemia that 3818

resulted in an increased lipid peroxidation, oxidative stress, and depressed metabolism 3819

of endogenous and exo. Res. Autism Spectr. Disord. 4, 119–155. 3820

https://doi.org/10.1016/j.rasd.2009.09.011 3821

Prandovszky, E., Gaskell, E., Martin, H., Dubey, J.P., Webster, J.P., McConkey, G. a., 2011. 3822

The neurotropic parasite Toxoplasma gondii increases dopamine metabolism. PLoS 3823

One 6. https://doi.org/10.1371/journal.pone.0023866 3824

Priddel, D., Carlile, N., Wheeler, R., 2000. Eradication of European rabbits (Oryctolagus 3825

cuniculus) from Cabbage Tree Island, NSW, Australia, to protect the breeding habitat of 3826

Gould’s petrel (Pterodroma leucoptera leucoptera). Biol. Conserv. 94, 115–125. 3827

Priddel, D., Carlile, N., Wilkinson, I., Wheeler, R., 2011. Eradication of exotic mammals from 3828

offshore islands in New South Wales, Australia, in: Veitch, C.R., Clout, M.N., Towns, D.R. 3829

(Eds.), Island Invasives: Eradication and Management. IUCN, Gland, Switzerland, pp. 3830

337–344. 3831

Proudfoot, G.A., Honeycutt, R.L., Slack, R.D., Ingraldi, M.F., 2006. Variation in DNA 3832

microsatellites of the ferruginous pygmy-owl (Glaucidium brasilianum). Conserv. Genet. 3833

7, 945–956. https://doi.org/10.1007/s10592-006-9138-5 3834

Puan, C.L., Baxter, G.S., Goldizen, A.W., Zakaria, M., Hafidzi, M.N., 2012. Barn owl predatory 3835

behavior and response to prey abundance: Towards an ecologically-based agricultural 3836

practice. Ornis Mongolica 1, 63–66. 3837

Purcell, K., Verner, J., Oring, L., 1997. A comparison of the breeding ecology of birds nesting 3838

176

in boxes and tree cavities. Auk 114, 646–656. 3839

Rattner, B.A., Horak, K.E., Lazarus, R.S., Goldade, D. a., Johnston, J.J., 2014a. Toxicokinetics 3840

and coagulopathy threshold of the rodenticide diphacinone in eastern screech-owls 3841

(Megascops asio). Environ. Toxicol. Chem. 33, 74–81. https://doi.org/10.1002/etc.2390 3842

Rattner, B.A., Lazarus, R.S., Elliott, J.E., Shore, R.F., van den Brink, N., 2014b. Adverse 3843

Outcome Pathway and Risks of Anticoagulant Rodenticides to Predatory Wildlife. 3844

Environ. Sci. Technol. 48, 8433–8445. https://doi.org/10.1021/es501740n 3845

Reece, R.L., Scott, P.C., Forsyth, W.M., Gould, J.A., Barr, D.A., 1985. Toxicity episodes 3846

involving agricultural chemicals and other substances in birds in Victoria, Australia. Vet. 3847

Rec. 525–527. https://doi.org/10.1136/vr.117.20.525 3848

Reith, D.M., Pitt, W.R., Hockey, R., 2001. Childhood poisoning in Queensland : An analysis of 3849

presentation and admission rates. J. Paediatr. Child Health 37, 446–450. 3850

Remington, J.S., Cavanaugh, E.N., 1965. Isolation of the encysted form of Toxoplasma gondii 3851

from human skeletal muscle and brain. N. Engl. J. Med. 273, 1308–1310. 3852

https://doi.org/10.1056/NEJM196512092732404 3853

Resch, H., Pietschmann, P., Krexner, E., Willvonseder, R., 1991. Decreased peripheral bone 3854

mineral content in patients under anticoagulant therapy with phenprocoumon. Eur. 3855

Heart J. 12, 439–441. 3856

Richomme, C., Afonso, E., Tolon, V., Ducrot, C., Halos, L., Alliot, A., Perret, C., Thomas, M., 3857

Boireau, P., Gilot-Fromont, E., 2010. Seroprevalence and factors associated with 3858

Toxoplasma gondii infection in wild boar (Sus scrofa) in a Mediterranean island. 3859

Epidemiol. Infect. 138, 1257–1266. https://doi.org/10.1017/S0950268810000117 3860

Ricketts, T.H., 2001. The matrix matters: effective isolation in fragmented landscapes. Am. 3861

Nat. 158, 87–99. https://doi.org/10.1086/320863 3862

Riegert, J., Fainová, D., Bystřická, D., 2010. Genetic variability, body characteristics and 3863

reproductive parameters of neighbouring rural and urban common kestrel (Falco 3864

tinnuculus) populations. Popul. Ecol. 52, 73–79. https://doi.org/10.1007/s10144-009-3865

0168-y 3866

177

Riley, S.P.D., Bromley, C., Poppenga, R.H., Uzal, F.A., Whited, L., Sauvajot, R.M., 2007. 3867

Anticoagulant Exposure and Notoedric Mange in Bobcats and Mountain Lions in Urban 3868

Southern California. J. Wildl. Manage. 71, 1874–1884. https://doi.org/10.2193/2005-3869

615 3870

Robertson, I., Leggoe, M., Dorling, P., Shaw, S., Clark, W., 1992. A retrospective study of 3871

poisoning cases in dogs and cats: comparison between a rural and an urban practice. 3872

Aust. Vet. J. 69, 194–195. 3873

Robinson, G.R., Holt, R.D., Gaines, M.S., Hamburg, S.P., Johnson, M.L., Fitch, H.S., Martinko, 3874

E.A., 1992. Diverse and contrasting effects of habitat fragmentation. Science (80-. ). 3875

257, 524–526. https://doi.org/10.1126/science.257.5069.524 3876

Robinson, M.H., Wheeler, S.H., 1983. A Radiotracking Study of Four Poisoning Techniques 3877

for Control of the European Rabbit, Oryctolagus cuniculus (L.). Aust. Wildl. Res. 10, 3878

513–520. 3879

Rodríguez-Estrella, R., Donázar, J.A., Hiraldo, F., 1998. Raptors as Indicators of 3880

Environmental Change in the Scrub Habitat of Baja California Sur, Mexico. Conserv. 3881

Biol. 12, 921–925. https://doi.org/10.1111/j.1523-1739.1998.97044.x 3882

Roll, U., Feldman, A., Novosolov, M., Allison, A., Bauer, A.M., Bernard, R., Böhm, M., Castro-3883

Herrera, F., Chirio, L., Collen, B., Colli, G.R., Dabool, L., Das, I., Doan, T.M., Grismer, L.L., 3884

2017. The global distribution of tetrapods reveals a need for targeted reptile 3885

conservation. Nat. Ecol. Evol. 1, 1677–1682. https://doi.org/10.1038/s41559-017-0332-3886

2 3887

Rosier, J.R., Ronan, N.A., Rosenberg, D.K., 2006. Post-Breeding Dispersal of Burrowing Owls 3888

in an Extensive California Grassland. Am. Midl. Nat. 155, 162–167. 3889

https://doi.org/10.1674/0003-0031(2006)155[0162:PDOBOI]2.0.CO;2 3890

Rueda, D., Campbell, K.J., Fisher, P., Cunninghame, F., Ponder, J.B., 2016. Biologically 3891

significant residual persistence of brodifacoum in reptiles following invasive rodent 3892

eradication, Galapagos Islands, Ecuador. Conserv. Evid. 13, 38. 3893

Ruiz-Suárez, N., Henríquez-Hernández, L.A., F. Valerón, P., Boada, L.D.B., Zumbado, M., 3894

Camacho, M., Almeida-González, M., Luzardo, O.P., 2014. Assessment of anticoagulant 3895

178

rodenticide exposure in six raptor species from the Canary Islands (Spain). Sci. Total 3896

Environ. 485, 371–376. 3897

https://doi.org/http://dx.doi.org/10.1016/j.scitotenv.2014.03.094 3898

Rutkowski, R., Rejt, L., Szczuka, A., 2006. Analysis of microsatellite polymorphism and 3899

genetic differentiation in urban and rural kestrels Falco tinnunculus (L.). Polish J. Ecol. 3900

54, 473–480. 3901

Salim, H., Noor, H.M.H.H., Hamid, N.H., Omar, D., Kasim, A., Abidin, C.M.R.Z., Hafidz, M.N., 3902

Noor, H.M.H.H., Omar, D., Kasim, A., Abidin, C.M.R.Z., 2015. The Effects of Rodenticide 3903

Residues Deposited in Eggs of Tyto alba to Eggshell Thickness The Effects of 3904

Rodenticide Residues Deposited in Eggs. Sains Malaysiana 44, 559–564. 3905

https://doi.org/10.17576/jsm-2015-4404-10 3906

Sánchez-Barbudo, I.S., Camarero, P.R., Mateo, R., 2012. Primary and secondary poisoning by 3907

anticoagulant rodenticides of non-target animals in Spain. Sci. Total Environ. 420, 280–3908

288. https://doi.org/10.1016/j.scitotenv.2012.01.028 3909

Saunders, D.A., 1989. Changes in the avifauna of a region, district and remnant as a result of 3910

fragmentation of native vegetation: the Wheatbelt of Western Australia. A case study. 3911

Biol. Conserv. 50, 99–135. 3912

Saunders, D.A., Doley, A., 2017. Culling farm wildlife for conservation and production on 3913

“Koobabbie”, a cereal and sheep growing property , in the northern wheatbelt of 3914

Western Australia. Aust. Zool. 1–15. https://doi.org/DOI: 3915

https://doi.org/10.7882/AZ.2017.046 3916

Saunders, D.A., Hobbs, R.J., Margules, C.R., 1991a. Biological Consequences of Ecosystem 3917

Fragementation: A review. Conserv. Biol. 5, 18–32. https://doi.org/10.1111/j.1523-3918

1739.1991.tb00384.x 3919

Saunders, D.A., Hobbs, R.J., Margules, C.R., 1991b. Biological Consequences of Ecosystem 3920

Fragmentation: A Review. Conserv. Biol. 5, 18–32. https://doi.org/10.1111/j.1523-3921

1739.1991.tb00384.x 3922

Saunders, D.A., Mawson, P.R., Dawson, R., 2014. Use of tree hollows by Carnaby’s Cockatoo 3923

and the fate of large hollow-bearing trees at Coomallo Creek, Western Australia 1969-3924

179

2013. Biol. Conserv. 177, 185–193. https://doi.org/10.1016/j.biocon.2014.07.002 3925

Saunders, D.A., Smith, G.T., Rowley, I., 1982. The availability and dimensions of tree hollows 3926

that provide nest sites for cockatoos (Psittaciformes) in Western Australia. Wildl. Res. 3927

9, 541–556. https://doi.org/10.1071/WR9820541 3928

Saunders, G., Kay, B., Parker, B., 1990. Evaluation of a Warfarin Poisoning Programme for 3929

Feral Pigs (Sus scrofa). Aust. Wildl. Res. 17, 525–533. 3930

Saunders, G.R., 1983. Evaluation of mouse-plague control techniques in irrigated sunflower 3931

crops. Crop Prot. 2, 437–445. https://doi.org/10.1016/0261-2194(83)90064-9 3932

Saunders, Ingram, J.A., 1995. Birds of southwestern Australia: an atlas of changes in 3933

distribution and abundance of the wheatbelt fauna, Birds of southwestern Australia: an 3934

atlas of changes in distribution and abundance of the wheatbelt fauna. Surrey Beattey 3935

& Sons Pty. Limited, Chipping Norton, NSW. 3936

Scelza, B.A., Bird, D.W., Bird, R.B., 2017. Bush Tucker, Shop Tucker: Production, 3937

Consumption, and Diet at an Aboriginal Outstation. Ecol. Food Nutr. 53, 98–117. 3938

https://doi.org/10.1080/03670244.2013.772513 3939

Schulze, L., 1966. Nest predation by Boobook Owl. South Aust. Ornithol. 24, 110. 3940

Sedgwick, E.H., Morrison, P.C., 1948. Observations on the lower Murchison RAOU Camp, 3941

September 1948. Emu 48, 212–242. 3942

Serieys, L.E.K., Armenta, T.C., Moriarty, J.G., Boydston, E.E., Lyren, L.M., Poppenga, R.H., 3943

Crooks, K.R., Wayne, R.K., Riley, S.P.D., 2015. Anticoagulant rodenticides in urban 3944

bobcats: exposure, risk factors and potential effects based on a 16-year study. 3945

Ecotoxicology 24, 844–862. https://doi.org/10.1007/s10646-015-1429-5 3946

Serieys, L.E.K., Lea, A.J., Epeldegui, M., Armenta, T.C., Moriarty, J., VandeWoude, S., Carver, 3947

S., Foley, J., Wayne, R.K., Riley, S.P.D., Uittenbogaart, C.H., 2018. Urbanization and 3948

anticoagulant poisons promote immune dysfunction in bobcats. Proc. R. Soc. B — Biol. 3949

Sci. 285, 20172533. 3950

Serventy, D.L., 1948. The Birds of the Swan River District, Western. Australia. Emu 47, 241–3951

286. https://doi.org/10.1071/MU947241 3952

180

Serventy, V., 1980. To Kill a Kookaburra. State Wildl. Auth. News Serv. 5–6. 3953

Shanahan, D.F., Possingham, H.P., Riginos, C., 2011. Models based on individual level 3954

movement predict spatial patterns of genetic relatedness for two Australian forest 3955

birds. Landsc. Ecol. 26, 137–148. https://doi.org/10.1007/s10980-010-9542-6 3956

Shore, R.F., Walker, L.A., Potter, E.D., Pereira, G., 2016. Second generation anticoagulant 3957

rodenticide residues in barn owls 2015. Lancaster, UK. 3958

Simon, J.A., Pradel, R., Aubert, D., Geers, R., Villena, I., Poulle, M.L., 2018. A multi-event 3959

capture-recapture analysis of Toxoplasma gondii seroconversion dynamics in farm 3960

cats. Parasites and Vectors 11, 1–13. https://doi.org/10.1186/s13071-018-2834-4 3961

Sims, V., Evans, K.L., Newson, S.E., Tratalos, J.A., Gaston, K.J., 2008. Avian assemblage 3962

structure and domestic cat densities in urban environments. Divers. Distrib. 14, 387–3963

399. https://doi.org/10.1111/j.1472-4642.2007.00444.x 3964

Sisk, T.D., Haddad, N.M., Ehrlich, P.R., 1997. Bird assemblages in patchy woodlands: 3965

Modeling the effects of edge and matrix habitats. Ecol. Appl. 7, 1170–1180. 3966

https://doi.org/10.1890/1051-0761(1997)007[1170:BAIPWM]2.0.CO;2 3967

Soderquist, T., Gibbons, D., 2007. Home-range of the Powerful Owl (Ninox strenua) in dry 3968

sclerophyll forest. Emu 107, 177–184. 3969

Soulé, M.E., Simberloff, D., 1986. What do genetics and ecology tell us about the design of 3970

nature reserves? Biol. Conserv. https://doi.org/10.1016/0006-3207(86)90025-X 3971

Speck, F.G., 1941. Gourds of the southeastern Indians. New Engl. Gourd Soc. 40–41. 3972

Sroka, J., Cencek, T., Ziomko, I., Karamon, J., Zwoliński, J., 2008. Preliminary assessment of 3973

ELISA, MAT, and LAT for detecting Toxoplasma gondii antibodies in pigs. Bull. Vet. Inst. 3974

Pulawy 52, 545–549. 3975

Stansley, W., Cummings, M., Vudathala, D., Murphy, L.A., 2014. Anticoagulant Rodenticides 3976

in Red-Tailed Hawks, Buteo jamaicensis, and Great Horned Owls, Bubo virginianus, 3977

from New Jersey, USA, 2008 – 2010. Bull. Environ. Contam. Toxicol. 92, 6–9. 3978

https://doi.org/10.1007/s00128-013-1135-z 3979

Stephenson, B.M., Minot, E.O., Armstrong, D.P., 1999. Fate of moreporks (Ninox 3980

181

novaeseelandiae) during a pest control operation on Mokoia Island, Lake Rotorua, 3981

North Island, New Zealand. N. Z. J. Ecol. 23, 233–240. 3982

Stone, W.B., Okoniewski, J.C., Stedelin, J.R., 2003. Anticoagulant rodenticides and raptors: 3983

Recent findings from New York, 1998-2001. Bull. Environ. Contam. Toxicol. 70, 34–40. 3984

https://doi.org/10.1007/s00128-002-0152-0 3985

Stone, W.B., Okoniewski, J.C., Stedelin, J.R., 1999. Poisoning of wildlife with anticoagulant 3986

rodenticides in New York. J. Wildl. Dis. 35, 187–193. https://doi.org/10.7589/0090-3987

3558-35.2.187 3988

Storr, G.M., Johnstone, R.E., 1988. Birds of the Swan Coastal Plain and adjacent seas and 3989

islands. Western Australian Museum, Perth. 3990

Stranger, R., 2003. The Birds of the Perth District, Western Australia. Robert Stranger, Perth. 3991

Suri, J., Sumasgutner, P., Hellard, E., Koeslag, A., Amar, A., 2016. Stability in prey abundance 3992

may buffer Black Sparrowhawks Accipiter melanoleucus from health impacts of 3993

urbanization. Ibis (Lond. 1859). 159, 38–54. 3994

Tasmania Parks and Wildlife Sevice, 2014. 2014 Evaluation Report: Macquarie Island Pest 3995

Eradication Project. Hobart, Tasmania. 3996

Taylor, M., Canberra Ornitholgists Group, 1992. Birds of the Australian Capital Territory: An 3997

Atlas. The Group and the National Capital Planning Authority, Canberra. 3998

Tella, J.L., Hiraldo, F., Donazar-Sancho, J.A., Negro, J.J., 1996. Costs and Benefits of Urban 3999

Nesting in the Lesser Kestrel, Raptors in Human Landscapes: Adaptations to Built and 4000

Cultivated Environments. Academic Press, New York. 4001

Thomas, M.O.W.G., Kutt, A.S., 1997. Owl populations and habitat: factors that could impact 4002

populations of native owls in the sugarcane growing areas in Queensland, Australian 4003

Centre for Tropical Freshwater Research, James Cook University of North Queensland. 4004

Townsville. https://doi.org/10.1017/CBO9781107415324.004 4005

Thomas, P.J., Mineau, P., Shore, R.F., Champoux, L., Martin, P. a., Wilson, L.K., Fitzgerald, G., 4006

Elliott, J.E., 2011. Second generation anticoagulant rodenticides in predatory birds: 4007

Probabilistic characterisation of toxic liver concentrations and implications for 4008

182

predatory bird populations in Canada. Environ. Int. 37, 914–920. 4009

https://doi.org/10.1016/j.envint.2011.03.010 4010

Torrey, E.F., Bartko, J.J., Lun, Z.R., Yolken, R.H., 2007. Antibodies to Toxoplasma gondii in 4011

patients with schizophrenia: a meta-analysis. Schizophr. Bull. 33, 729–736. 4012

https://doi.org/sbl050 [pii]\r10.1093/schbul/sbl050 4013

Tosh, D.G., Shore, R.F., Jess, S., Withers, A., Bearhop, S., Ian Montgomery, W., McDonald, 4014

R.A., 2011. User behaviour, best practice and the risks of non-target exposure 4015

associated with anticoagulant rodenticide use. J. Environ. Manage. 92, 1503–1508. 4016

https://doi.org/10.1016/j.jenvman.2010.12.014 4017

Trejo-Macías, G., Estrada, A., Cabrera, M., 2007. Survey of helminth parasites in populations 4018

of Alouatta palliata mexicana and A. pigra in continuous and in fragmented habitat in 4019

southern Mexico. Int. J. Primatol. 28, 931–945. 4020

Trost, S., Olsen, J., 2016. Common Mynas attack Common Brushtail Possums and nesting 4021

Southern Boobook, and diet of the nesting Boobooks. Aust. F. Ornithol. 33, 22–24. 4022

https://doi.org/10.20938/afo33022024 4023

Trost, S., Olsen, J., Rose, A.B., Debus, S.J.S., 2008. Winter diet of Southern Boobooks Ninox 4024

novaeseelandiae in Canberra 1997-2005. Corella 32, 66–70. 4025

Twigg, L.E., Lowe, T.J., Martin, G.R., Gray, G.S., 1999. A Review of the Anticoagulant 4026

Pesticide Pindone. 4027

Twigg, L.E., Singleton, G.R., Kay, B.J., 1991. Evaluation of Bromadiolone Against House 4028

Mouse (Mus domesticus) Populations in Irrigated Soybean Crops. I . Efficacy of Control. 4029

Wildl. Res. 18, 265–274. 4030

Vidal, D., Alzaga, V., Luque-Larena, J.J., Mateo, R., Arroyo, L., Viñuela, J., 2009. Possible 4031

interaction between a rodenticide treatment and a pathogen in common vole 4032

(Microtus arvalis) during a population peak. Sci. Total Environ. 408, 267–271. 4033

https://doi.org/10.1016/j.scitotenv.2009.10.001 4034

Walker, L.A., Chaplow, J.S., Llewellyn, N.R., Pereira, M.G., Potter, E.D., Sainsbury, A.W., 4035

Shore, R.F., 2011. Anticoagulant rodenticides in predatory birds 2011: a Predatory Bird 4036

183

Monitoring Scheme (PBMS) report. Lancaster, UK. 4037

Walker, L.A., Turk, A., Long, S.M., Wienburg, C.L., Best, J., Shore, R.F., 2008. Second 4038

generation anticoagulant rodenticides in tawny owls (Strix aluco) from Great Britain. 4039

Sci. Total Environ. 392, 93–98. https://doi.org/10.1016/j.scitotenv.2007.10.061 4040

Walters, J., 1991. Application Of Ecological Principles To The Management Of Endangered 4041

Species: The Case Of The Red-Cockaded Woodpecker. Annu. Rev. Ecol. Syst. 22, 505–4042

523. https://doi.org/10.1146/annurev.ecolsys.22.1.505 4043

Wan, H.Y., Cushman, S.A., Ganey, J.L., 2018. Habitat fragmentation reduces genetic diversity 4044

and connectivity of the Mexican spotted owl: A simulation study using empirical 4045

resistance models. Genes (Basel). 9, 1–21. https://doi.org/10.3390/genes9080403 4046

Wang, J., 2018. Effects of sampling close relatives on some elementary population genetics 4047

analyses. Mol. Ecol. Resour. 18, 41–54. https://doi.org/10.1111/ijlh.12426 4048

Ward, D.J., 2008. An ecological assessment of secondary poisoning risk in the Australian 4049

sugarcane industry. Queensland University of Technology. 4050

Warner, R., 1968. The role of introduced diseases in the extinction of the endemic Hawaiian 4051

avifauna. Condor 70, 101–120. 4052

Watson, J.E.M., Whittaker, R.J., Freudenberger, D., 2005. Bird community responses to 4053

habitat fragmentation: How consistent are they across landscapes? J. Biogeogr. 32, 4054

1353–1370. https://doi.org/10.1111/j.1365-2699.2005.01256.x 4055

Weaving, M.J., White, J.G., Isaac, B., Cooke, R., 2011. The distribution of three nocturnal bird 4056

species across a suburban - forest gradient. Emu 111, 52–58. 4057

https://doi.org/10.1071/MU09117 4058

Wedding, C.J., Weihong, J., Brunton, D.H., 2010. Implications of visitations by Shore Skinks 4059

Oligosoma smithi to bait stations containing brodifacoum in a dune system in New 4060

Zealand. Pacific Conserv. Biol. 16, 86–91. 4061

Weidensaul, C.S., Colvin, B.A., Brinker, D.F., Huy, J.S., 2011. Use of Ultraviolet Light as an Aid 4062

in Age Classification of Owls. Wilson J. Ornithol. 123, 373–377. 4063

https://doi.org/10.1676/09-125.1 4064

184

Weir, S.M., Yu, S., Talent, L.G., Maul, J.D., Andreson, T.A., Salice, C.J., 2015. Improving reptile 4065

ecological risk assessment : Oral and dermal toxicity of pesticides to a common lizard 4066

species (Sceloporus occidentalis). Environ. Toxicol. Chem. 34, 1778–1786. 4067

https://doi.org/10.1002/etc.2975 4068

Wells, B., 2007. Boobooks breeding in Perth. Boobook 25, 14. 4069

Western Australian Museum, n.d. Feral Bees | Western Australian Museum [WWW 4070

Document]. URL http://museum.wa.gov.au/explore/online-exhibitions/cockatoo-4071

care/feral-bees (accessed 6.3.15). 4072

Whitford, K.R., 2002. Hollows in jarrah (Eucalyptus marginata) and marri (Corymbia 4073

calophylla) trees: I. Hollow sizes, tree attributes and ages. For. Ecol. Manage. 160, 201–4074

214. https://doi.org/10.1016/S0378-1127(01)00446-7 4075

Williams, N.S.G., Morgan, J.W., Mccarthy, M. a, Mark, J., 2006. Local Extinction of Grassland 4076

Plants: The Landscape Matrix Is More Important than Patch Attributes. Ecology 87, 4077

3000–3006. 4078

Woinarski, J.C.Z., Burbidge, A.A., Harrison, P.L., 2015. Ongoing unraveling of a continental 4079

fauna: Decline and extinction of Australian mammals since European settlement. Proc. 4080

Natl. Acad. Sci. 112, 4531–4540. https://doi.org/10.1073/pnas.1417301112 4081

Wolf, A., Cowen, D., Paige, B., 1939. Human toxoplasmosis: occurrence in infants as an 4082

encephalomyelitis verification by transmission to animals. Science (80-. ). 89, 226–227. 4083

https://doi.org/10.1126/science.89.2306.226 4084

Young, J., De Lai, L., 1997. Population declines of predatory birds coincident with the 4085

introduction of Klerat rodenticide in north Queensland. Aust. Bird Watch. 17, 160–167. 4086

Yu, L., Shen, J., Su, C., Sundermann, C. a., 2013. Genetic characterization of Toxoplasma 4087

gondii in wildlife from Alabama, USA. Parasitol. Res. 112, 1333–1336. 4088

https://doi.org/10.1007/s00436-012-3187-0 4089

4090

4091

185

Co-author Statements 4092

4093

Signed co-author statements verifying my role in the production of papers and manuscripts which 4094

make up chapters in this thesis are provided in this section. 4095

Chapter 2 4096

4097

186

Chapter 4 4098

4099

187

Chapter 5 4100

4101

188

Chapter 6 4102

4103

189

Copies of original publications 4104

I include below copies of the first page of published peer-reviewed journal articles corresponding to 4105 chapters in this thesis. No licenses are required to reproduce these papers either in part or in full 4106 when included as part of a PhD thesis per the Elsevier license agreement: 4107 https://service.elsevier.com/app/answers/detail/a_id/565/track/AvMKOAoHDv8W~QaHGnwa~yKg_4108 38qZS75Mv9z~zj~PP_6/ 4109

Chapter 2 4110

4111

190

Chapter 3 4112

4113