Post on 06-Nov-2021
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
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“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”
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Abstract The effects of habitat fragmentation on native wildlife can vary depending on the
type of land use occurring in the matrix between remaining habitat fragments. I used
Australian boobooks (Ninox boobook) in Western Australia to investigate interactions
between matrix type and four different potential threatening processes: secondary
poisoning by anticoagulant rodenticides (ARs); limitation of juvenile dispersal and impacts
on spatial genetic structure; breeding site availability; and infection by the parasite
Toxoplasma gondii.
I also conducted a literature review on the use and regulation of ARs in Australia and
published accounts of non-target impacts in order to contextualise exposure patterns
observed in boobooks. The review revealed records of confirmed or suspected poisoning
across 37 vertebrate species in Australia. World literature relating to AR exposure in
reptiles suggests that they may be less susceptible to AR poisoning than birds and mammals.
This relative resistance may create unevaluated risks for wildlife and humans in Australia
where reptiles are more abundant than in cooler regions where AR exposure has been
studied in greater depth.
I analysed AR residues in boobook livers across multiple habitat types. Second
generation anticoagulant rodenticides were detected in 72.6% of individuals sampled. Total
AR concentration correlated positively with the proportion of urban land use within an area
approximately the size of a boobook’s home range centred on the point where the sample
was collected. ARs originating in urban habitat probably pose a substantial threat to
boobooks and other predatory wildlife species.
No spatial genetic structure was evident in boobooks across habitat types. I
observed one individual dispersing at least 26km from its natal home range across urban
habitat. The apparent permeability of anthropogenically altered landscapes probably
explains the lack of spatial genetic structure and is likely related to the observed ability of
boobooks to use resources in both urban and agricultural matrices.
Boobooks did not appear to be limited by the availability of suitable nesting sites in
urban or agricultural landscapes. Occupancy did not change significantly over the duration
of the study in remnants provided with artificial nest boxes in either landscape type.
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However, in one instance, boobooks successfully used a nest box located in an urban
bushland. Nest boxes may be a useful management tool in highly-altered areas where
natural hollows are unavailable.
Toxoplasma gondii seropositivity in boobooks did not vary significantly by landscape
type but was more prevalent in individuals sampled during cooler wetter times of year. Risk
of exposure due to greater cat abundance in urban and agricultural landscapes may be
offset by creation of environmental conditions less favourable to the survival of T. gondii
oocysts in soil.
Taken together, this body of research demonstrates variation in relationships
between different types of habitat fragmentation and threatening processes related to
fragmentation. This research also raises questions about how habitat fragmentation is
discussed and studied in the context of species which are capable of making extensive use
of matrix habitat. I recommend greater consideration of the concept of “usable space”
when studying fragmentation impacts in habitat generalists.
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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
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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
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dissertation is a testament to the quality of his photography and the mileage I have gotten
out of his photos of my work. I sincerely hope I can repay my debt as he continues his PhD
and I look forward to future and ongoing collaborations.
I wish to express my sincere thanks to Dr. Jamie Tedeschi for her patience and
expertise in introducing me to the world of genetic analysis and to Louise Pallant and
A/Prof. Annette Koenders for their advice and assistance on serological testing. Training a
field ecologist to do lab work is surely a painful experience and I am grateful that they
attempted it.
I particularly appreciate Jerry Olsen contributing data from his boobook banding
projects as well as helpful advice and friendly correspondence throughout my PhD.
I also thank Ben Jones and Yvonne Sitko for helping me to communicate the results
of my work to the public. Without their efforts, much of my work would not have made it
to the people who can actually use it.
I especially thank Rachele Bernasconi, Casper Avenant, Melissa Karlinski, Emily Lette,
Rosh McCallum, and Charlie Phelps for their moral support and for tolerating my
eccentricity, frightening desktop, and questionable musical taste through the writing
process. Your contributions to my sanity were critical to getting this dissertation finished.
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Statement of contribution of others
Research Funding
The Holsworth Wildlife Research Endowment via The Ecological Society of Australia
BirdLife Australia Stuart Leslie Bird Research Award
Edith Cowan University School of Science Postgraduate Student Support Award
Eastern Metropolitan Regional Council’s Healthy Wildlife Healthy Lives program
The Society for the Preservation of Raptors
Sian Mawson
Stipend
Edith Cowan University Postgraduate Research Scholarship
Edith Cowan University Merit Award
Supervision
Dr. Robert A. Davis
Dr. Allan H. Burbidge
Field Assistance
Casper Avenant, Rachele Bernasconi, Simon Cherriman, Jakeb Cumming, Angela Febey, Sian
Glazier, Melissa Hetherington, Tyson Isles, Michael Just, Candice Le Roux, Gabe Mach, Paul
Radley, Calan Rance, Geoffrey Schoonakker, Nakisa Shahrestani, Lia Smith, Steven Spragg,
Paula Strickland, Mitch Wright
Laboratory Technical Assistance and Advice
A/Prof. Annette Koenders, Louise Pallant, Dr. Jamie Tedeschi,
Co-Authors
Dr. Janet Anthony, Dr. Allan H. Burbidge, Simon Cherriman, Dr. Robert A. Davis, Dr. Siegfried
Krauss, Dr. Cheryl A. Lohr, A/Prof. Peter B. S. Spencer
The research included in this dissertation is my original work. I conceived and
developed all hypotheses, led all field work, designed or conducted the majority of
analysis,wrote all first drafts, and made the majority of edits to subsequent drafts. The co-
authors listed above contributed to one or more chapters in at least one of the following
ways: advice on experimental design, assistance in fieldwork, data analysis, and editing of
drafts. I am the lead author on all published articles and manuscripts. My roles in each
chapter are detailed in the “Co-author Statements” section.
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Publications arising from this research
I am submitting this thesis as a thesis by publication. Chapters 2 and 3 are
reformatted versions of the published journal articles. A single reference list is provided for
the entire thesis following the final chapter. No permission is needed to reproduce these
articles as part of a PhD thesis. The first pages of published chapters 2 and 3 can be found in
the section entitled “Copies of original publications”
Chapter 2
Lohr, M. T., and R. A. Davis (2018). Anticoagulant rodenticide use, non-target impacts and
regulation: A case study from Australia. Science of the Total Environment. 634:1372–
1384.
Chapter 3
Lohr, M. T. (2018). Anticoagulant rodenticide exposure in an Australian predatory bird
increases with proximity to developed habitat. Science of the Total Environment.
643:134–144.
Chapter 4
Lohr, M. T., P. B. S. Spencer, S. Krauss, J. Anthony, A. H. Burbidge, and R. A. Davis.
Widespread genetic connectivity in Australia’s most common owl, despite extensive
habitat fragmentation. The Condor: Ornithological Applications. (In Preparation).
Chapter 5
Lohr, M. T., S. Cherriman, A. H. Burbidge, and R. A. Davis. Artificial nest box supplementation
does not affect Australian boobook (Ninox boobook) occupancy in fragmented habitats
in south-western Australia. Wildlife Research. (In Review).
Chapter 6
Lohr, M. T., C. A. Lohr, A. H. Burbidge, and R. A. Davis. Toxoplasma gondii seropositivity
across urban and agicultural landscapes in an Australian owl. Veterinary Parasitology.
(In Preparation).
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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
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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
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Chapter 5 Artificial nest box supplementation does not affect Australian boobook (Ninox boobook)
occupancy in fragmented habitats in south-western Australia ................................................... 100
Abstract ........................................................................................................................................... 100
Introduction .................................................................................................................................... 101
Nest Competition and Predation ................................................................................................ 102
Impacts of Nest Boxes in Conservation ...................................................................................... 103
Knowledge Gaps.......................................................................................................................... 104
Methods .......................................................................................................................................... 106
Study Sites ................................................................................................................................... 106
Surveys ........................................................................................................................................ 107
Nest box construction and placement ........................................................................................ 108
Nest Box Monitoring ................................................................................................................... 111
Statistical Analysis ....................................................................................................................... 112
Results ............................................................................................................................................. 112
Discussion........................................................................................................................................ 114
Surveys ........................................................................................................................................ 114
Nest Box Use ............................................................................................................................... 115
Conclusion ................................................................................................................................... 119
Acknowledgments ........................................................................................................................... 119
Chapter 6 Toxoplasma gondii seropositivity across urban and agricultural landscapes in an
Australian owl ......................................................................................................................... 120
Abstract ........................................................................................................................................... 120
Introduction .................................................................................................................................... 121
Effects of Toxoplasma gondii on Humans and Wildlife .............................................................. 122
Predatory Birds and Toxoplasma gondii Infection ...................................................................... 123
Aims............................................................................................................................................. 124
Methods .......................................................................................................................................... 125
Sample Collection ....................................................................................................................... 125
Serological Testing ...................................................................................................................... 126
Statistical Analysis ....................................................................................................................... 127
Results ............................................................................................................................................. 128
Discussion........................................................................................................................................ 131
Landscape Type ........................................................................................................................... 133
Age .............................................................................................................................................. 133
Injury Status ................................................................................................................................ 134
Season ......................................................................................................................................... 134
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Anticoagulant Rodenticide Exposure .......................................................................................... 135
Acknowledgments ........................................................................................................................... 136
Chapter 7 Summary, Synthesis, and Management Implications ................................................. 137
Summary of major findings: ............................................................................................................ 137
Objective 1. Critically review literature on anticoagulant rodenticide exposure in native wildlife
in Australia to clarify its role as a threatening process. .............................................................. 137
Objective 2. Investigate the relationship between exposure to anticoagulant rodenticides and
urban and agricultural fragmentation. ....................................................................................... 138
Objective 3. Determine if urban and agricultural fragmentation influence boobook genetic
structure. ..................................................................................................................................... 138
Objective 4. Examine whether nest box supplementation increases site occupancy at
unoccupied sites and whether this effect differs between urban and agricultural landscapes. 139
Objective 5. Explore patterns of Toxoplasma gondii seropositivity in boobooks across the urban,
agricultural, and natural landscapes. .......................................................................................... 139
Synthesis ......................................................................................................................................... 140
Management Recommendations ................................................................................................... 144
Anticoagulant Rodenticides ........................................................................................................ 144
Nest Box Supplementation ......................................................................................................... 145
References ............................................................................................................................... 146
Co-author Statements .............................................................................................................. 185
Copies of original publications .................................................................................................. 189
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List of Figures Figure 3.1 Percentages of Southern Boobooks (n=73) in Western Australia exposed to rodenticides
stratified by total rodenticide liver concentration (mg/kg) thresholds indicating potential outcomes.
.............................................................................................................................................................. 54
Figure 3.2 Percentages of Southern Boobooks (n = 73) exposed to multiple anticoagulant
rodenticides in Western Australia. ....................................................................................................... 55
Figure 3.3 Mean total anticoagulant rodenticide concentration (mg/kg) in liver tissue of Southern
Boobooks (n= 71) in Western Australia by season. .............................................................................. 56
Figure 4.1 Sample locations of genotyped Australian Boobooks (Ninox boobook) in Western
Australia. (“metro” = urban and suburban areas of Perth represented by squares, “rural” = forested
area surrounding the Perth Metropolitan area represented by an “x” , “Southwest WA” = forested
areas to the south of Perth represented by triangles, “Wheatbelt” = highly-fragmented agricultural
landscapes represented by crosses, and “other” = Goldfields and Pilbara regions, represented by
black circles, ‘other’ = Goldfields and Pilbara regions of Western Australia). ...................................... 77
Figure 4.2 A corellogram showing genetic correlation values (r) as a function of distance (kms) using
eight microsatellite markers in a subset of Australian Boobooks (Ninox boobook) n=98 from the
Perth metropolitan area, adjacent exurban areas and the Perth Hills. U and L are 95% confidence
intervals around the null hypothesis of no spatial genetic structure. No significant genetic structure
is shown at any distance class. ............................................................................................................. 83
Figure 4.3 Principal coordinate analysis results based on eight microsatellite loci in Australian
Boobooks (Ninox boobook) in Western Australia. Clustering does not correspond to potential
populations and is driven by two common alleles and their heterozygotes at the locus Nst15. Blue =
161/161, Green = 161/uncommon allele, Purple = 163/161, Orange = 163/uncommon allele, Red =
163/163, Black = no result. ................................................................................................................... 83
Figure 4.4 Principal coordinate analysis results based on seven microsatellite loci (i.e. no Nst15 – see
Fig 3) in Australian Boobooks in Western Australia. No clustering is apparent across or within six
sampled regions (“Exurbs” = areas immediately surrounding but not within the Perth Metropolitan
area, “Perth Hills” = an area of continuous forest east of Perth, “Perth Metro” = urban and suburban
areas of Perth, ‘Remote WA’ = Goldfields and Pilbara regions of Western Australia, “Southwest WA”
= forested areas to the south of Perth, “Wheatbelt” = highly-fragmented agricultural landscapes
existing primarily between the “Remote” region and all other regions). ............................................ 85
Figure 4.5 Visualization of Australian Boobooks (Ninox boobook) sampled from six regions in
Western Australia (“Exurbs” = areas immediately surrounding but not within the Perth Metropolitan
area, “Perth Hills” = an area of continuous forest east of Perth, “Perth Metro” = urban and suburban
areas of Perth, ‘Remote WA’ = Goldfields and Pilbara regions of Western Australia, “Southwest WA”
= forested areas to the south of Perth, “Wheatbelt” = highly-fragmented agricultural landscapes
existing primarily between the “Remote” region and all other regions) using the STRUCTURE results
from CLUMPAK comparing number of inferred genetic clusters (K) from 1-6. The data support a
single genetic cluster. Each line represents an individual. The proportion of colours in each line
represents the proportion of membership of each individual in each cluster. .................................... 86
Figure 4.6 Plot of Evanno et al.’s (2005) delta K (ΔK) based on inferred genetic clusters (populations)
ranging from 2 to 5 in Australian Boobooks (Ninox boobook) sampled from Western Australia. ....... 87
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Figure 5.1 Locations of survey sites in in southwestern Western Australia: urban landscapes in the
Perth Metropolitan Area, continuous bushland in the Perth Hills, and agricultural landscapes within a
60km radius of Kellerberrin, Western Australia. ................................................................................ 107
Figure 5.2 Attachment system used to hang nest boxes used in this study. ...................................... 110
Figure 5.3 A nest box installed in one of the remnant bushlands in an agricultural landscape in
Western Australia. .............................................................................................................................. 111
Figure 6.1 Seasonal Toxoplasma gondii seroprevalence in Australian Boobooks (Ninox boobook) in
Western Australia. Width of the bars is representative of sample size. ............................................ 130
Figure 6.2 Toxoplasma gondii seroprevalence in meat juice from deceased Australian Boobooks
(Ninox boobook) in Western Australia in four different categories of anticoagulant rodenticide
exposure (A= ≤ 0.01 mg/kg, B=0.01 mg/kg – 0.10 mg/kg, C 0.10 mg/kg - 0.50mg/kg, D ≥ 0.50mg/kg)
Width of the bars is representative of sample size. ........................................................................... 131
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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
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Co-author Statements 4092
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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
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Chapter 4 4098
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Chapter 5 4100
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Chapter 6 4102
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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
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Chapter 3 4112
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