The Ecological Value of Suburban Golf Courses in South-East Queensland
Author Hodgkison, Simon Charles
Published 2006
Thesis Type Thesis (PhD Doctorate)
School Centre for Innovative Conservation Strategies
DOI https://doi.org/10.25904/1912/304
Copyright Statement The author owns the copyright in this thesis, unless stated otherwise.
Downloaded from http://hdl.handle.net/10072/367634
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The ecological value of suburban golf courses in south-east Queensland
by
SIMON CHARLES HODGKISON BSC MSC
Centre for Innovative Conservation Strategies Griffith University Gold Coast Campus PMB50 Gold Coast Mail Centre Bundall QLD 9726 Australia
Thesis submitted for the degree of Doctor of Philosophy
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Statement of Originality
The work presented in this thesis has never been previously submitted for a degree or diploma in any university and contains no material previously written or published by another person, unless otherwise acknowledged.
December 1, 2005 Simon Charles Hodgkison
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iv ACKNOWLEDGEMENTS
This study could not have been completed without the support and assistance of a considerable number of people to whom I am extremely grateful. I’d first like to thank my principle supervisor, Jean-Marc Hero for providing ongoing support and encouragement and for doing his utmost to keep me on the ecological straight and narrow with critical feedback on data analysis and draft chapters. Special thanks also to Marc for continuing to read chapters while technically on paternity “leave” (and to Narinder and Priya for allowing him to do so!). Your help has been greatly appreciated. I’d also like to thank my associate supervisors: Jan Warnken for support above and beyond the call of duty, for helping me view the project from an urban planning perspective, for providing valuable advice on GIS and for pre-PhD employment on the turkeyfest; and Daryl Jones for all your help during the confirmation and early development stages of the project.
Thanks to Griffith University for funding this project through a Postgraduate Research Scholarship and to the Queensland Parks and Wildlife Service for allowing access to study sites.
Many others provided valuable assistance over the course of the project. I’d particularly like to thank Michael Arthur for his unusual (but very helpful) brand of statistical advice. Many thanks also to Clyde Wild for providing great support particularly in the early stages of the project and for eagerly offering statistical and ecological advice even when approached at inopportune moments. Sincere thanks also to Carla Catterall, Harry Hines, Damian White, Luke Shoo, Naomi Doak, Steve Phillips, Donna Hazell, Monique Van Sluys, Guy Castley, Ed Meyer, Michaela Guest, Brian Zancola, Cuong Tran, Kerry Kriger, Abraham Mijares, Chris Gregory and everyone in the ecology lab group for valuable feedback and impromptu discussions. Every bit has helped.
This study was made entirely possible by the staff and management of the California Creek Golf Course, Gailes Golf Club, Gold Coast Country Club, Gold Coast-Burleigh Golf Club, Gainsborough Greens Golf Club, Helensvale Golf Club, Keperra Country Club, McLeod’s Golf Club, Oxley Golf Club, Parkwood International Golf Club, The Pacific Golf Club, The Grand Golf Club, Redland Bay Golf Club, Southport Golf Club, Surfers Paradise Golf Club, St Lucia Golf Links, Tweed Heads-Coolangatta Golf Club, Virginia Golf Club and Wynnum Golf Club. All clubs were enthusiastically involved with the project from the beginning and allowed me unlimited access to sites and golf carts. Their positive attitude to the project and the
v ACKNOWLEDGEMENTS
potential for environmental improvement was a source of inspiration during tougher moments and confirmed in my mind that the conservation potential of the industry was worth investigating. I’d particularly like to thank Jon Pemberthy, Rod Cook, Jeff Gambin, Colin Gibbs, Andrew Baker, Justin Kelly, Barry Lemke, Graham Simms, Kerry Cahill, Harley Kruse, Ian Farnfield, Dennis Newcombe, Don Gregory, Charlie James, David Doyle, Greg Plummer, Eric Martin, Peter Daly, Sean and Vicki Keeley and Peter Lonergan for offering valuable information on golf management practices and local wildlife habits. Thanks also to Joellen Zeh and Larry Woolbright of Audubon International and Bob Taylor from the Sports Turf Research Institute for information on golf course ecology.
Thanks to Jutta Masterton and Andrew Bryant for much appreciated assistance with field equipment. Thanks also to Keith Wilson from DNR for help with Mapview, to the Tweed Shire Council for providing rectified aerial images of the Coolangatta-Tweed Heads Golf Club and Mariola Hoffmann from Nathan campus for DCDB and remnant veg layers.
I’d like to thank everyone who’s helped make the last few years so enjoyable; all the postgrads and academics in EAS, especially Daniel and Michelle Stock, Michaela Guest and Andrew Melville, Pascal Scherer, Nicole Thornton, Naomi Doak, Luke Shoo, Skye Page, Kristy Morris, Jason (does this look normal) Vandermerwe, Olaf Meynecke, Andrew Growcock, Rodney Duffey, Liz West, the Citric Hens, Colin Hutchins, Shosh Fogelman, Jo Oakes, Michael Brickhill, Kerry Kriger, Chris Gregory, Ruth Young and James Webley. Thanks also to Chris, Foz, JB, Pete, Howard Moon, Vince Noir and Bob Fossil.
Finally I’d like to thank my family: Mum and Dad, Jim, Susan, Keira and Emily for being an ongoing source of support (and charity food packages). Without sounding like a bad episode of Webster, I love you all very much. There’s no way this could have been completed without you. Finally I’d like to thank my grandma Mandy for being one of the greatest influences in my life. I will always feel lucky to have known you. This thesis is dedicated to you.
vi ABSTRACT
Information is required on the ecological value of all urban land types in order to provide an ecological basis for urban zoning decisions, to predict development impacts and to identify much needed opportunities for off-reserve conservation. One land type (the suburban golf course) has experienced a dramatic global proliferation in recent decades, as courses are increasingly constructed as part of new housing developments and resorts. Golf courses account for a growing proportion of the urban land area in Australia and will have an increasingly significant impact on urban biodiversity. The nature of their influence is however contentious. While many suggest golf courses have a negative impact on biodiversity, others believe they can provide refugial habitat for native wildlife. This potential refuge value has been nurtured by the golf industry as a way to improve its environmental reputation. However while the industry has initiated programs to enhance the quality of habitats on golf courses, it is uncertain whether such small-scale conservation efforts can have more than a cosmetic effect. The ubiquity of suburban golf courses makes their possible ecological contribution more significant and thus worthy of investigation.
This study assessed the conservation status of suburban golf courses in south-east Queensland Australia between 2001-2004, by comparing assemblages of birds, reptiles, mammals and amphibians on 20 representative golf courses with those in 10 nearby eucalypt remnants and with bird assemblages in 10 suburban areas. The ecological characteristics of wildlife utilising golf courses were compared with those shared by animals common to residential areas and to native eucalypt forest. Local changes in bird diversity were assessed following the clearance of small vegetation remnants on a suburban golf course. Finally, the ecological value gained by increasing the size and complexity of native habitats on golf courses was assessed, by investigating the extent to which differences in biodiversity among golf courses were attributed to the size, shape and complexity of local, landscape and regional habitats on and adjacent to golf courses.
vii ABSTRACT
Golf courses displayed extreme variation in conservation value. While a number of golf courses had significant refuge value, supporting high densities of regionally threatened vertebrates, most supported only common urban-adapted species and therefore failed to realise that potential. Wildlife assemblages in residential areas and on most golf courses were more homogenised than those in eucalypt forests and were generally dominated by species with broad ecological tolerances. In contrast, ecological specialists were restricted to eucalypt forests and a minority of golf courses. The clearance of even small remnants of native vegetation on a single golf course had a significant homogenising effect on local bird assemblages with a diverse range of regionally threatened birds being replaced by a small number of urban- adapted species.
Golf courses that did have refuge value had the capacity to accommodate most regionally threatened species. In general, golf courses were a better refuge for threatened birds and mammals than for threatened reptiles and amphibians. The relative absence of threatened herpetofauna may reflect heightened sensitivity to habitat isolation, faster rates of local decline, to increased local threats (i.e. predation or herbicide exposure) or a difference in the extent to which their habitats have been compromised.
Differences in biodiversity among golf courses were attributed to environmental factors acting at local, landscape and regional scales. While the local diversity of all vertebrates was partly determined (and therefore restricted by) regional influences, the local abundance and species richness of threatened vertebrates still closely reflected the size and structural complexity of on-course habitats. The diversity of all vertebrates increased with the area of native vegetation retained locally. Species-area curves were observed among reptiles, birds and mammals. Contrary to other studies, there was no distinct threshold in the species-area relationship. Threatened species gradually disappeared from the landscape as patch sizes decreased below 5ha. The lack of any distinct threshold highlights the dangers of proposing spatially explicit
viii ABSTRACT
guidelines to ecologically sound development. Patches should always be as large as possible if they are intended to provide refuge to threatened vertebrates. Remnant size was however, co-correlated with structural complexity, with the understorey of smaller remnants often cleared to increase course playability. Biodiversity in small remnants (<5ha) may therefore be increased by enhancing the structural complexity of local habitats. Bird diversity increased with foliage height diversity and native grass cover. Mammal diversity increased with tree density, native grass cover and the abundance of hollows. Reptile diversity increased with the abundance of woody debris and declined with the proportion of turfgrass cover. The diversity of amphibians increased with waterbody diversity, the complexity of aquatic and riparian vegetation, the number of connecting streams and declined with the steepness and proportion of turfgrass cover on waterbody banks.
Golf courses can evidently provide locally valuable refuges for threatened vertebrates and have a clear opportunity to make an important localised contribution to urban wildlife conservation in SEQ. The extent to which this potential is realised will however depend on the extent to which ecological criteria are incorporated into golf course design and management practices. The current low conservation status of most existing golf courses reflects a historic lack of regulation and formal Environmental Impact Assessment within the golf industry. This has seen few golf courses retain any substantial area of complex core vegetation.
While the industry is aware of its conservation potential, the economic pressures affecting modern course designs will tend to restrict the area of native vegetation that can be retained (particularly among housing development courses where there is pressure to maximise the area of land available for housing). Legislation recognising and protecting the value of small remnants in new golf developments may be required if the golf industry is to realise its conservation potential.
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x TABLE OF CONTENTS PAGE NO.
Statement of Originality iii Acknowledgements v Abstract vii Table of Contents xi List of Tables xv List of Figures xvii List of Appendices xix
CHAPTER 1 GENERAL INTRODUCTION 1.1. The Ecological Impacts of Urbanisation 1 1.2. Traditional Approach to Urban Wildlife Conservation 1 1.3. The Need for a Hierarchical Approach to Conservation 2 1.4. The Importance of the Landscape Matrix 3 1.5. The Need for an Opportunistic Approach to Conservation 3 1.6. The Global Proliferation of Golf Courses 4 1.7. The Ecological Potential of Golf Courses 5 1.8. Positive Changes in the Golf Industry 6 1.9. Research on the Ecological Value of Golf Courses 6 1.10. The Significance of Golf Courses in South-east Queensland 8 1.11. Thesis Outline 9
CHAPTER 2 THE CONSERVATION VALUE OF SUBURBAN GOLF COURSES 11 2.1. Introduction 11 2.2. Aims 11 2.3. Methods 12 2.3.1. Site Selection 12 2.3.2. Sub-site Selection 14 2.3.3. Fauna Surveys - Birds 14 - Reptiles 14 - Mammals 14 - Amphibians 15 2.4. Data Analysis 15 2.4.1. Biodiversity Estimates 15 2.4.2. Species Land Preferences 15 2.4.3. Assemblage Composition 15 2.4.4. Refuge Value 16 2.5. Results 17 2.5.1. Comparing Biodiversity Estimates 17 2.5.2. Local Species Abundances 18 2.5.3. Assemblage Composition 20 2.5.4. Suburban ‘Exploiting, Avoiding and Tolerating’ Species 26 2.5.5. Refuge Value 26 2.6. Discussion 29 2.6.1. Implications for New Golf Course Construction 30 2.6.2. Refuge Value of Golf Courses 30 2.6.3. The Ecological Role of Habitat on Golf Courses 32
xi CHAPTER 3 ECOLOGICAL CHARACTERISTICS OF WINNERS AND LOSERS 35 3.1. Introduction 35 3.2. Aims 36 3.3. Methods 36 3.4. Data Analysis 37 3.5. Results 37 Birds – Microhabitat Selection 37 – Body Size 38 – Diet and Foraging Strategy 38 – Migrant Status 39 – Territoriality and Aggression 39 Amphibians 39 Mammals 40 Reptiles 43 3.6. Discussion 43 3.6.1. Birds 43 3.6.2. Mammals 45 3.6.3. Reptiles 46 3.6.4. Amphibians 47 3.6.5. Characteristics of Wildlife on Golf Courses 48
CHAPTER 4 A REVIEW OF LANDSCAPE ECOLOGY THEORY 49 Introduction 49 The Theory of Island Biogeography Theory 49 The Theory of Metapopulation Dynamics 49 History of Landscape Ecology Research 50 Patch Size 50 Patch Shape 51 Habitat Complexity 51 Landscape Connectivity 52 Landscape Context 53 Importance of the Matrix 53 The Effects of Fragmentation 54 The Importance of Spatial Scale 54 The Importance of Temporal Scale 55 Conclusion 55
CHAPTER 5 ASSESSMENT OF HABITAT ON GOLF COURSES 57 5.1. Introduction 57 5.2. Methods 58 5.2.1. Local Scale Environmental Factors 58 5.2.2. Landscape Environmental Factors 59 5.2.3. Regional Environmental Factors 59 5.3. Results 61 5.3.1. Terrestrial Habitats 61 5.3.2. Aquatic Habitats 62 5.4. Discussion 72
xii CHAPTER 6 FACTORS INFLUENCING BIODIVERSITY ON GOLF COURSES 73 6.1. Introduction 73 6.2. Methods 74 6.2.1. Univariate Analyses 74 6.2.2. Multivariate Analyses 76 6.3. Results 77 6.3.1. Birds 77 6.3.2. Mammals 78 6.3.3. Reptiles 79 6.3.4. Amphibians 80 6.3.5. Multivariate 89 6.4. Discussion 90 6.4.1. Regional Influences 90 6.4.2. Landscape Influences 91 6.4.3. Local Patch Scale Influences 93
CHAPTER 7 SPECIES-AREA RELATIONSHIPS ON GOLF COURSES 97 7.1. Introduction 97 7.2. Methods 98 7.2.1. Data Collection 98 7.2.2. Minimum Patch Occupancy and Occupancy Rates 98 7.2.3. Changes in the Composition of Avoiders, Tolerators and Exploiters 98 7.3. Results 98 7.3.1. Minimum Patch Occupancy 98 Birds 98 Reptiles 100 Mammals 100 7.3.2. Patch Occupancy Rates 102 7.3.3. Change in Assemblage Composition 102 7.4. Discussion 106 7.4.1. The Species-Area Accumulation Curve 108 7.4.2. Patch Occupancy by Individual Species 106 7.4.3. Changes in Assemblage Composition 108 7.4.4. Management Implications 110
CHAPTER 8 CHANGE IN THE LOCAL BIRD ASSEMBLAGE FOLLOWING HABITAT LOSS ON A GOLF COURSE 113 8.1. Introduction 113 8.2. Methods 113 8.2.1. Biotic and Abiotic Surveys 113 8.2.2. Data Analysis 113 8.3. Results 114 8.3.1. Change in Biodiversity 114 8.3.2. Species Responses 114 8.3.3. Species Composition 119 8.4. Discussion 119
xiii CHAPTER 9 MANAGEMENT RECOMMENDATIONS 121 9.1. Introduction 121 9.2. Local Scale Actions – Implications for Golf Course Superintendents 121 9.2.1. Increase Vertical Complexity of Vegetation 122 9.2.2. Increase Mean Tree Density 122 9.2.3. Retain Dead Trees and Hollow-bearing Trees 124 9.2.4. Increase the Proportion of Native Grass Cover 124 9.2.5. Retain Logs, Fallen Wood and Rotting Woody Debris 124 9.2.6. Increase Aquatic Vegetation Complexity 126 9.2.7. Decrease Turfgrass Cover 126 9.2.8. Naturalise Waterbody Designs 126 9.2.9. Avoid stocking Waterbodies with Fish 128 9.2.10. Aquatic Weed Management 128 9.2.11. Factors that could Deter Conservation Efforts 129 - Increased Fire Risk 129 - Danger of Tree Fall 130 - Cost of Planting 132 - Member Backlash 132 9.3. Landscape Scale Actions – Implications for Golf Course Architects 132 9.3.1. Maximise the Area of Native Vegetation Retained 133 9.3.2. Maximise Individual Patch Sizes 133 9.3.3. Increase Patch Width 133 9.3.4. Maximise Vegetation Connectivity on the Course 134 9.3.5. Maximise the Diversity of Local Waterbody Types 134 9.3.6. Factors Limiting the Use of Ecological Design Criteria 134 - Economic Constraints of Housing Development Courses 136 - Aesthetic Expectations of Resort Courses 136 - Hydrological/Engineering Constraints 136 9.3.7. Using ‘Natural’ Course Designs for Economic Gain 137 9.4. Regional Scale Actions – Implications for Urban Planners 137 9.4.1. Bushland Areas should not be Targeted for Development 137 9.4.2. Research Required to Direct Ecologically Sound Zoning Decisions 138 9.4.3. Legislation to Control the Ecological Value of Golf Courses 138
REFERENCES 140 APPENDICES 172
xiv LIST OF TABLES PAGE NO.
Table 2.1 Sites surveyed on the Gold Coast and Brisbane 12 Table 2.2a Total abundance of terrestrial bird species in each land type 21 Table 2.2b Total abundance of terrestrial and aquatic birds in each land type 22 Table 2.3 Total abundance of reptile, amphibian and mammal species in each land type 23 Table 6.1 Final abiotic variables used in multiple regression analysis 75 Table 6.2 Significant variables from multiple regression between local, landscape and regional environmental variables and the site abundance and species richness of birds, reptiles, mammals and amphibians 82 Table 6.3 Significant variables from multiple regression between local, landscape and regional environmental variables and the site abundance and species richness of threatened birds, reptiles, mammals and amphibians 83 Table 6.4 Significant variables from multivariate regression between environmental characteristics measured at local, landscape and regional scales and the abundance of specific ecological groups of birds, reptiles, mammals and amphibians 88 Table 6.5 Results of BIO-ENV multivariate correlation between the local assemblage of birds, reptiles, mammals and amphibians and environmental characteristics 89
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xvi LIST OF FIGURES PAGE NO.
Figure 2.1 Map of south-east Queensland showing the regional distribution of study sites 13 Figure 2.2 Mean site abundance and species richness of all birds, aquatic birds, terrestrial birds, reptiles, mammals 19 Figure 2.3 SSH-MDS Ordination plot showing the relative similarity in species composition among sites for each vertebrate group 27 Figure 2.4 The refuge value of individual sites indicated by the mean site abundance of threatened birds, reptiles, mammals and amphibians 28 Figure 3.1 Total site abundance of birds with different ecological characteristics in residential areas, golf courses and eucalypt forests 40 Figure 3.2 Total site abundance of amphibians with different ecological characteristics on golf courses and eucalypt forests 41 Figure 3.3 Total site abundance of different mammal groups 42 Figure 3.4 Total site abundance of different reptile groups 42 Figure 5.1 Schematic diagrams showing the methods used to measure landscape and regional indices from rectified aerial photographs in MapInfo 60 Figure 5.2 A) The total area of edge and core habitat, water and turfgrass on each golf course B) the size of all habitat patches and C) the area of core habitat on and adjacent to each golf course in a radius of 200m and 500m 63 Figure 5.3 Relationship between golf course size and A) patch size, B) total area of vegetation on the course and C) proportion of the course devoted to native vegetation 64 Figure 5.4 Mean structural complexity of local habitats at each site 65 Figure 5.5 Plots showing the relationship between patch width and the structural complexity of local habitats 66 Figure 5.6 Aquatic environmental site variables A) Area of water on each course, B) Mean proportion of vegetation cover, C) Mean proportion of waterbody bank fringed by reeds 67 Figure 5.7 Schematic maps of golf courses G1-G5 68 Figure 5.8 Schematic maps of golf courses G6-G10 69 Figure 5.9 Schematic maps of golf courses G11-G15 70 Figure 5.10 Schematic maps of golf courses G16-G20 71 Figure 6.1 Partial plots for univariate multiple regression between bird diversity indices and combined environmental variables 84 Figure 6.2 Partial plots for univariate multiple regression between mammal diversity indices and combined environmental variables 85 Figure 6.3 Partial plots for univariate multiple regression between reptile diversity indices and combined environmental variables 86 Figure 6.4 Partial plots for univariate multiple regression between amphibian diversity indices and combined environmental variables 87 Figure 7.1 Minimum patch sizes occupied by individual bird species 99 Figure 7.2 Minimum patch sizes occupied by individual reptile species 101 Figure 7.3 Minimum patch sizes occupied by individual mammal species 103 Figure 7.4 Species-area accumulation curves for birds, reptiles and mammals based on mean minimum patch sizes occupied 104
xvii Figure 7.5 Change in density of suburban avoiders, tolerators and exploiters and the species richness of suburban-avoiders with patch size 105 Figure 8.1 Schematic map of golf course G17 before and after housing development 115 Figure 8.2 Change in site abundance of suburban-exploiting, tolerating and avoiding birds on golf course G17 before, during and after development 116 Figure 8.3 Changes in the abundance of bird species before, during and after an on-course housing development at golf course G17 117 Figure 8.4 SSH-MDS Ordination plot showing the change in bird assemblage composition on golf course G17 relative to other golf courses suburban residential areas and eucalypt remnants 118 Figure 9.1 Increasing the vertical complexity of native vegetation will increase bird, reptile and mammal diversity 123 Figure 9.2 Increasing tree density will increase mammal diversity 123 Figure 9.3 Retaining dead and hollow-bearing trees will increase habitat availability for arboreal mammals and some birds and reptiles 125 Figure 9.4 Retaining native grass in out-of-play areas will provide habitat for mammals, birds and reptiles 125 Figure 9.5 Woody debris will support greater reptile diversity if retained in situ 125 Figure 9.6 Natural waterbodies that retain structurally diverse aquatic and riparian vegetation will support greater amphibian diversity than artificial waterbodies with little vegetation 127 Figure 9.7 Stocking waterbodies with fish will reduce frog diversity 131 Figure 9.8 Aquatic weed infestations can be minimised by aerating water and restricting nutrient run-off 131 Figure 9.9 Controlled burning is required for fuel reduction and biodiversity management 131 Figure 9.10 Maximising the area of vegetation in out-of-play areas will increase bird, reptile, mammal and amphibian diversity 135 Figure 9.11 Maximising patch width will increase reptile diversity 135 Figure 9.12 Semi-permanent and ephemeral waterbodies support a higher diversity of threatened amphibians than permanent ponds 135
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Appendix 1 Classifying reptile, mammal and amphibian species into urban avoiding, tolerating and exploiting species based on the opinion af six local ecologists. 172 Appendix 2 Ecological characteristics of terrestrial bird species 173 Appendix 3 Ecological characteristics of terrestrial bird species (continued) 174 Appendix 4 Ecological characteristics of individual reptile species 175 Appendix 5 Ecological characteristics of individual amphibian species 176 Apendix 6 Ecological characteristics of individual mammal species 177 Appendix 7 Occupancy rates of bird species in different patch size categories 178 Appendix 7 Occupancy rates of bird species in different patch size categories 179 Appendix 8 Occupancy rates of reptile species in different patch size categories 180 Appendix 9 Occupancy rates of mammal species in different patch size categories 180
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xx CHAPTER 1 GENERAL INTRODUCTION
1.1 THE ECOLOGICAL IMPACTS OF URBANISATION
Urban land types have become the subject of targeted ecological research as efforts are made to understand and ameliorate the environmental impacts of urbanisation (Gilbert, 1991). Urbanisation is globally recognised as a significant threat to biodiversity, responsible for the localised decline and extinction of many native wildlife species (Wilson, 1986; Adams, 1994; Marzluff and Ewing, 2001). Wildlife are principally affected by habitat loss and fragmentation resulting from urban land clearing (Usher, 1987; Collinge, 1996; Fahrig, 1999; Marzluff and Ewing, 2001). Secondary factors further degrade the quality of remnant habitats by disrupting nutrient and hydrological cycles, restricting access to resources, obstructing wildlife movements and increasing levels of pollution, disturbance, noise, predation and competition (Brittingham and Temple, 1983; Robinson and Wilcove, 1994; Knight and Gutzwiller, 1995). Limiting the impacts of urban development has become a high conservation priority. Most developed nations therefore employ some form of Environmental Impact Assessment (EIA) to minimise development impact and urban planning strategies to protect the ecological integrity of urban landscapes (Benkendorff, 2000). However, while these processes are in place, there is often insufficient information on the ecological consequences of small-scale landscape change to make accurate EIA and urban planning decisions (Thompson et al., 1997; Benkendorff, 2000). 1.2 TRADITIONAL APPROACH TO URBAN WILDLIFE CONSERVATION Until recently, smaller habitat remnants within the urban landscape have received relatively little research attention from ecologists (Franklin, 1993; Jules, 1997; Vandermeer, 1997; Blair and Launer, 1997; Freeman, 1999). Traditional approaches to urban wildlife conservation focussed primarily on the retention of large habitat reserves (Franklin, 1993; Fischer and Lindenmayer, 2002a). An accumulated body of ecological theory, derived from the principles of island biogeography and species-area relationships demonstrate that habitat value increases with reserve size (MacArthur and Wilson, 1967; Levins, 1969; Diamond, 1975). Conservation strategies have therefore invariably stressed the need to protect large reserves that retain areas of quality ‘interior’ habitat (Wilson and
1 Willis, 1975; Terborgh, 1975; Margules et al., 1982; Noss, 1983; Blake and Karr, 1987; Patterson, 1987; Doak and Mills, 1994). 1.3 THE NEED FOR A HIERARCHICAL APPROACH TO CONSERVATION While the need to protect large habitat reserves is unquestionable and must form the basis of urban conservation strategies, it has often been misinterpreted to suggest smaller habitat fragments are without ecological value (Franklin, 1993; Fischer and Lindenmayer, 2002a). In reality, the extent of urban habitat loss is such that no opportunity for conservation can afford to be overlooked. Small-intermediate remnants represent the vast majority of conservation opportunities in the urban landscape (Primack, 1993). While small remnants can be edge-effected and therefore of limited value to interior-dependent wildlife, they are not without ecological value (Barrett et al., 1994; Fischer and Lindenmayer, 2002a). Wildlife respond to the landscape at different scales (Wiens, 1989; Kotliar and Wiens, 1990). Some species are less vulnerable to habitat fragmentation than others (Terborgh, 1986; Laurance, 1991; Barrett et al., 1994; Andrén et al., 1997; Mönkkönnen and Reunanen, 1999; Fischer and Lindenmayer, 2002a) and can therefore benefit from the protection of small habitat remnants. Many ecologists therefore recommend a hierarchical approach to conservation that combines efforts at a range of scales on a variety of land-types (Holling, 1992; Hostetler, 1999; Hostetler and Holling, 2000; Savard et al., 2000). While small habitat remnants have restricted independent value, they can complement reserve networks by providing additional resources, softening edge-effects and increasing habitat connectivity (Grover and Slater, 1994; Bentley and Catterall, 1997; Semlitsch and Bodie, 1998; Gilfedder and Kirkpatrick, 1998; Fischer and Lindenmayer, 2002a). Conservation efforts on small land types can also have cumulative benefits at a regional level (Savard et al., 2000). The possible contribution of smaller remnants should therefore not be underestimated (Schwartz and van Mantgem, 1997; Marzluff and Ewing, 2001; Fischer and Lindenmayer, 2002a; 2002b). At the same time, the inherent ecological limitations of small remnants need to be clearly understood to ensure small-scale conservation efforts have realistic goals and target appropriate species. Research is required to assess the role that can be played by small remnants retained in parks, paddocks, gardens and golf courses, their capacity to assist metapopulation persistence and their role in the source-sink dynamics of fragmented landscapes (Franklin, 1993; Grover and Slater, 1994; Zuidema et al., 1996; Sisk et al., 1997; Semlitsch and Bodie, 1998; Gilfedder and Kirkpatrick, 1998; Fischer and Lindenmayer, 2002a).
2 1.4 THE IMPORTANCE OF THE LANDSCAPE MATRIX Another problem of traditional urban biodiversity strategies has been to ignore the influence of the urban matrix (Simberloff et al., 1992; Franklin, 1993; Beir and Noss, 1998). Many early ecological studies were based on a binary view that only distinguished areas of habitat and non-habitat (Lindenmayer et al., 2003). McIntyre and Barrett (1992) considered this approach overly simplistic and proposed the concept of a ‘variegated landscape’ that recognises possible gradients in habitat quality among land types. This approach is particularly relevant in heterogeneous urban environments. Wildlife assemblages in habitat remnants can be affected by the nature and juxtaposition of land types in the surrounding matrix (Harris, 1984; Laurance, 1990; 1991; 1994; Fahrig and Merriam, 1994; Malcolm, 1997; Gascon et al., 1999). The matrix can act as a selective filter, permitting the movement of some species, but restricting others (Laurance, 1997; Gascon et al., 1999). Depending on the nature and intensity of local and emanating threats, land types within the matrix can reduce productivity and survivorship in adjacent habitats (Janzen, 1986; Hutchings, 1991; Laurance, 1994; Tocher et al., 1997) or buffer reserves from external threatening processes (Laurance and Yensen, 1991; Mesquita et al., 1999). The urban matrix is therefore neither benign in its influence, nor uniformly inhospitable to wildlife. Urban conservation strategies must therefore consider not only the size and quality of habitat reserves, but the nature and juxtaposition of land types in the intervening urban matrix (Franklin, 1993). Information on the ecological value and influence of all urban land types is required to make ecologically responsible urban zoning decisions. 1.5 THE NEED FOR AN OPPORTUNISTIC APPROACH TO CONSERVATION Ultimately, the key impediment to urban wildlife conservation is a lack of opportunity to protect or restore habitat. Demand for access to urban land is intense. Ecological priorities must compete with social and commercial interests in a decision- making process that is primarily based on economic criteria (Freeman, 1999). A flexible unbiased approach is therefore required to identify all opportunities to retain habitat within the urban landscape (Catterall and Kingston, 1993). While conservation efforts should be made at a range of scales, the urban landscape is typically dominated by smaller land types. These are generally limited to providing habitat for animals that respond to the landscape at a fine scale. The real challenge is to retain habitat for species that respond to the landscape at a broad level; animals that require large areas of habitat or need to move substantial distances in the course of their daily and seasonal activities
3 (eg. large terrestrial mammals). These animals are particularly vulnerable to the fragmentation of their natural habitat and require a strategic, landscape-level approach to be adequately protected (Soulé, 1991). This demand has lead to a shift in urban planning policy in recent decades, whereby ecological priorities are incorporated throughout the planning process (Searns, 1995; Linehan et al., 1995). Attempts have been made to maintain and restore urban landscape integrity by creating networks of interconnected habitat, using a combination of formal reserves and off-reserve habitats on private land and in public ‘open space’ areas (Ahern, 1991; Searns, 1995; Linehan et al., 1995; Forman and Collinge, 1997). To achieve this goal, urban planners have therefore considered the capacity for open-space areas (eg. recreational parks, gardens, riverside ‘greenways’) to perform ecological as well as social or economic functions by retaining wildlife habitat (Linehan et al., 1995; Searns, 1995; Ndubisi et al., 1995; Johnson, 1995; Tilton, 1995; McGuckin and Brown, 1995; Greenway and Simpson, 1996; Sodhi et al., 1999; Pirnat, 2000; Viles and Rosier, 2001). These ‘multiple-use’ areas are not intended to form the basis of urban conservation strategies, however they may have some capacity to complement existing reserve networks by retaining additional habitat (Searns, 1995). While multiple-use areas have theoretical potential, it is important to define their ecological limitations and ensure there is a sound ecological basis to their use in urban planning strategies (Hobbs, 1997; Ehrenfeld, 2000). Research is required to assess the ecological value and limitations of off-reserve habitats in marginal land types. Urban biodiversity inventories have been conducted in recent decades (Kato et al., 1997; Matthews et al., 1988; Hermy and Cornelis, 2000; Mörtberg and Wallentinus, 2000; Nakamura and Short, 2001) recognising conservation value in linear ‘connector’ parks (Sodhi et al., 1999), storm-water mitigation areas and wastewater settling ponds (Greenway and Simpson, 1996). 1.6 THE GLOBAL PROLIFERATION OF GOLF COURSES The ecological value of one of the most ubiquitous suburban open-space land types (the suburban golf course) has become the subject of speculation in recent decades (Tatnall, 1991; Pearce, 1993). This interest has however been generated as much by concerns for their potentially adverse environmental impact as by their capacity to contribute to urban conservation strategies (Tatnall, 1991; Pearce, 1993). This speculation is in response to a recent global proliferation in the rate of golf course
4 construction, which has seen golf courses built at a rate of more than one per day in the USA and one per week in the UK throughout the 1980’s (Pearce, 1993). Since the early 1980’s, many parts of North America, Europe, Asia and Australia have experienced a boom in new golf course development, sponsored largely by foreign investment from the Asian ‘tiger’ economies (Chen, 1991; Pearce, 1993). Golf courses have economic value for the golf, real estate and tourism industries (Warnken et al., 2001). Golf courses increase property values (Schwanke, 1997; Warnken and Buckley, 1997) and broaden the recreation opportunity spectrum in tourism destinations (Priestly, 1995; Warnken et al., 2001). New golf course developments are therefore often funded by housing corporations and actively encouraged by local governments (Warnken et al., 2001). These factors have been a catalyst in the global proliferation of golf courses (Pearce, 1993). Golf courses account for a growing proportion of the urban land area in many parts of the world (Chen, 1991; Pearce, 1993). Australia currently has more than 500 suburban golf courses, accounting for as much as 5% of the urban land area in Australia’s capital cities (Dawson, 2000). That proportion is predicted to increase with many new golf developments planned (Warnken et al., 2001). Golf courses will therefore inevitably have some impact on urban biodiversity. The nature of their impact is however contentious. While some suggest golf courses have the capacity to act as an urban wildlife refuge (Hawthoorn, 1971; Wheeler, 1972; Maffei, 1978; Green and Marshall, 1987; Tatnall, 1991; Brennan, 1992; Tietge, 1992) others believe they have limited ecological value and have warned that the proliferation of golf courses may exacerbate the ecological impacts of urbanisation (Pleuramom, 1992; Pearce, 1993). 1.7 THE ECOLOGICAL POTENTIAL OF GOLF COURSES From an ecological perspective, there are reasons to suspect golf courses may have value for some threatened wildlife species. In the context of degraded suburban landscapes, golf courses represent relatively large areas of open space (60ha on average), in which there are few barriers to deter wildlife movement, relatively low levels of noise, traffic and disturbance and fenced protection from predators on some courses (Tietge, 1992). While playing areas (fairways, tees and greens) have limited value for wildlife (Moul and Elliott, 1994), most courses retain native vegetation in rough and out-of-play areas (Doak, 1992). These areas could provide wildlife habitat. At the same time, the environmental criticisms that have been levelled at the golf industry have a strong foundation. Concerns have specifically targeted the heavy use of
5 water, pesticides and chemical fertilisers, the use of exotic ornamental vegetation and the tendency to alter local topography during golf course construction (Tietge, 1992; Doak, 1992; Pearce, 1993; Platt, 1994; Smart et al., 1993; Hurdzan, 1996; Snow, 1996). These practices have generated a perception that golf courses have limited value for native wildlife and raised concerns that the recent proliferation of golf courses may have a negative impact on urban biodiversity (Pearce, 1993; Warnken et al., 2001). 1.8 POSITIVE CHANGES IN THE GOLF INDUSTRY Recent trends in the golf industry do however suggest that environmental standards are improving and that there may be some cause for optimism. In recent decades the use of pesticides and chemical fertilisers has been controlled by strict legislation (Tietge, 1992; Dobereiner, 1992; Schiffman, 1994). Golf course architects have moved toward more natural course designs that retain structurally complex, native vegetation in rough and out-of-play areas (Smart et al., 1993; Klemme, 1995; Perrett, 1996; Terman, 1997; Dawson, 2000; Mogford, 2000). Financial pressures have also lead many golf clubs to streamline their maintenance costs by allowing rough and out-of-play areas to revert to a more natural state, thereby reducing the area of land that is actively maintained (AGCSA, 1998; Smith, 1998). By designating areas as potential ‘wildlife habitat’, golf clubs have often inadvertently enhanced their reputation within the local community (Smith, 1998; AGCSA, 1998). Natural designs have proved popular with the golfing community and are therefore economically viable (Smart et al., 1993). Golf courses present a rare situation where economic/social circumstances favour the restoration of habitat, albeit on a small scale. Golf courses can readily implement restoration efforts given their strong community links and the presence of an existing workforce that is equipped to maintain the local environment. As such, golf courses represent an opportunity for small-scale, off-reserve wildlife conservation. While there will clearly be limits to the conservation outcomes that can be achieved on individual golf courses (given their modest size), their current ubiquity in suburban landscapes makes their regional influence (positive or negative) more substantial, and thus worthy of investigation. It is therefore important to determine the extent to which habitats on golf courses can support species that are threatened by urbanisation. 1.9 RESEARCH ON THE ECOLOGICAL VALUE OF GOLF COURSES A growing number of ecologists have investigated the capacity for suburban golf courses to provide refugial habitat for regionally threatened wildlife. Maffei (1978) found that an urban golf course in Massachusetts, USA, supported a range of songbirds,
6 waterbirds and a number of small ground mammals. Subsequent North American studies have shown that golf courses can support high songbird (Moul and Elliott, 1994; Jones et al., 2005) and waterbird species richness (White and Main, 2005). Others have demonstrated that golf courses can support endangered flora (Green and Marshall, 1987; Brennan, 1992) and threatened wildlife including the Big Cypress Fox squirrel (Jodice and Humphrey, 1992), Orange-bellied parrot (Ambrose, 1997), Red-headed woodpecker (Rodewald et al., 2005) and Burrowing owl (Smith et al., 2005). Recent research has investigated avian reproductive success on golf courses in the United States, with mixed results (LeClerc et al., 2005; Stanback and Seifert, 2005). Prior to the commencement of this study, very few researchers had assessed the relative conservation value of suburban golf courses by systematically comparing wildlife diversity on golf courses, with that found in adjacent land types. Terman (1997) had demonstrated that a well-vegetated golf course in Kansas, USA, supported bird species richness comparable to that in an adjacent natural prarie. Blair (1996) had found that suburban golf courses in California and Ohio supported intermediate bird species diversity, higher than that found in urban parks and residential areas, but lower than that in natural areas and recreational reserves. Another study in Phoenix, Arizona (Hostetler and Knowles-Yanez, 2003) had shown that bird species diversity varies dramatically among golf courses and other urban land types and that land-use was therefore a poor indicator of local habitat value. Research conducted in the northern hemisphere at the time of this study (unbeknownst to the author), demonstrated that golf courses can act as a refuge for birds and insects in rural England (Tanner and Grange, 2005) and for birds in desert regions of New Mexico, USA (Merola-Zwartjes and DeLong, 2005). Yet another study (LeClerc and Cristol, 2005), found that suburban golf courses in Virginia, USA were of little refuge value for birds of conservation significance and were of no greater value than residential or agricultural areas. The capacity for golf courses to provide habitat for threatened birds clearly varies regionally, depending on the surrounding landscape and the type of land displaced by golf courses. The refuge value of suburban golf courses is also likely to vary among taxonomic groups. To date, studies investigating the conservation value of suburban golf courses have all focussed on their capacity to provide habitat for birds. The extent to which golf courses can accommodate other less-mobile fauna is uncertain. Many golf courses are surrounded by residential land. This may restrict the capacity for wildlife to colonise or persist on isolated golf courses. An investigation of the ecological value of suburban golf
7 courses that contrasts the level of utilisation by a range of vertebrates would provide greater understanding of the potential ecological role that could be played by small off- reserve habitat remnants in urban landscapes. Research is also required to determine if the conservation value of suburban golf courses can be enhanced through improved golf course design and management practices. Golf industry manuals outline a range of management strategies that can be used to enhance wildlife diversity on golf courses (AGSCA, 1998; AGU, 1998). These are derived from landscape ecology principles and essentially promote increased size, structural complexity and heterogeneity of vegetation in rough and out-of-play areas. These principles are adapted from studies typically conducted at much larger scales. It is uncertain whether small-scale conservation efforts (such as those made on suburban golf courses) can benefit vertebrate species that are regionally threatened by urbanisation. This can only be determined by investigating abiotic-biotic relationships at a small scale. Studies conducted at the time of this project investigated factors influencing bird diversity on golf courses in the USA, producing varying results. While some found that bird species richness increases with the area of forest retained on golf courses (Jones et al., 2005; LeClerc and Cristol, 2005) or with the area and complexity of riparian vegetation (Merola-Zwartjes and DeLong, 2005), others have found that bird species richness on golf courses is determined predominantly by surrounding environmental characteristics and that local on-site variables are relatively unimportant (Porter et al., 2005). To date no-one has assessed the conservation value of Australian suburban golf courses, the capacity for suburban golf courses to provide habitat for wildlife other than birds, or considered factors influencing reptile, mammal and amphibian diversity on golf courses and bird diversity on Australian golf courses. 1.10 THE SIGNIFICANCE OF GOLF COURSES IN SOUTH-EAST QUEENSLAND This study investigates the ecological value of suburban golf courses in south-east Queensland (SEQ), Australia; a subtropical urban centre and tourism destination where golf courses occur in extremely high regional densities. An assessment of the relative ecological value of golf courses would have particular relevance in SEQ for several reasons. Firstly, the region (which includes Brisbane, the Sunshine Coast and the Gold Coast) is characterised by intense urban land-use conflict. SEQ currently accommodates Australia’s fastest rate of urban population growth (Graymore et al., 2002), and the nation’s second-highest level of regional biodiversity (Roberts, 1979; Pianka and Schall, 1981). Biological losses to urban development are therefore inevitable. Much of SEQ’s
8 lowland habitat has already been cleared (Catterall and Kingston, 1993) and remaining areas are threatened by ongoing housing development. All opportunities to retain lowland habitat must therefore be explored (Catterall and Kingston, 1993). In this context, the ecological contribution of marginal land types and open-space areas is expected to be important. Golf courses are the dominant open-space land type in SEQ. The region is one of Australia’s principal tourism destinations and golf courses are considered an integral part of the local tourism product. Local councils have therefore traditionally encouraged the construction of new golf courses. South-east Queensland is currently home to more than 130 eighteen-hole golf courses and another 39 have received planning approval on the Gold Coast alone (Warnken et al., 2001). As the dominant open-space land type in SEQ, golf courses have some ecological responsibility. There is however, no legislation governing the use of native vegetation on golf courses in SEQ. An evaluation of the ecological contribution of golf courses would be timely. This study assesses the current and potential value of golf courses as habitat for native vertebrate wildlife (birds, reptiles, mammals and frogs) and defines the ecological limits of their possible contribution to urban wildlife conservation. The study uses a landscape ecology approach to investigate the relationship between biodiversity and the structure and composition of landscape elements at a range of scales. This will identify opportunities to enhance the ecological value of suburban golf courses. 1.11 THESIS OUTLINE
CHAPTER 2: THE CONSERVATION VALUE OF SUBURBAN GOLF COURSES This chapter investigates the conservation value of golf courses in SEQ by comparing bird, reptile, mammal and amphibian assemblages on golf courses with those in nearby eucalypt remnants and with bird assemblages in adjacent suburban residential areas.
CHAPTER 3: THE ECOLOGICAL CHARACTERISTICS OF WINNERS AND LOSERS Chapter 3 identifies ecological characteristics shared by animals that persist in residential suburban areas, on golf courses and in eucalypt forest remnants. This will identify groups of species that are at greatest risk of extinction, those that require natural habitats to persist and those that can utilise modified habitats (eg. golf courses).
CHAPTER 4: REVIEW OF LANDSCAPE ECOLOGY THEORY Chapter 4 provides a review of existing landscape ecology theory, outlining factors that influence the distribution and abundance of wildlife species.
9 CHAPTER 5: ASSESSMENT OF VEGETATION ON GOLF COURSES Chapter 5 describes the size, spatial configuration and structural complexity of the environment on and adjacent to golf courses in SEQ.
CHAPTER 6: FACTORS INFLUENCING BIODIVERSITY ON GOLF COURSES Chapter 6 investigates environmental factors influencing local biodiversity on golf courses and determines the extent to which high biodiversity on some golf courses is attributed to local design and management practices.
CHAPTER 7: SPECIES-AREA RELATIONSHIPS ON GOLF COURSES Chapter 7 investigates the relationship between vegetation patch size and individual species, outlining the lower limits below which individual species do not persist. This will provide an understanding of the spatial requirements of different vertebrate groups, particularly those that are threatened by urbanisation.
CHAPTER 8: CHANGES IN THE LOCAL BIRD ASSEMBLAGE FOLLOWING HABITAT LOSS
ON A GOLF COURSE Many existing golf courses face development pressure. This chapter documents changes in the bird assemblage following the clearance of vegetation for housing.
CHAPTER 9: MANAGEMENT RECOMMENDATIONS Chapter 9 defines the role of course superintendents, course architects and urban planners in enhancing local biodiversity. This will ensure all efforts to improve biodiversity on golf courses have realistic goals and a clear understanding of the mechanisms required for ecological improvement.
10 CHAPTER 2
THE CONSERVATION VALUE OF SUBURBAN GOLF COURSES
2.1 INTRODUCTION A number of studies have suggested golf courses have conservation value simply because they support high species diversity (Maffei, 1978; Green and Marshall, 1992; Moul and Elliott, 1994). The use of diversity indices as a measure of habitat value has however been widely criticised because it provides no information on the conservation status or ecological significance of individual species (Peterken, 1974; Margules and Usher, 1981; Alatalo, 1981; Walker, 1992; Rossi and Kuitenen, 1996; Catterall et al., 1998; Magnusson, 2002). The concept of ‘representativeness’ (sensu Austin and Margules, 1984); i.e. the capacity to support a representative collection of species found within a given geographic region; is believed to more adequately reflect the motivation behind conservation efforts. Assessments of conservation value should therefore value land types that accommodate declining species above those that support high species diversity (Margules and Nicholls, 1987; Huxels and Hastings, 1999). Studies conducted in the USA have shown that suburban golf courses have variable conservation value, but that some do provide refuge to regionally threatened birds (Terman, 1997; Blair, 2001; Hostetler and Knowles-Yanez, 2003; Merola-Zwartjes and DeLong, 2005). The extent to which golf courses accommodate less mobile wildlife is uncertain. Many golf courses are surrounded by residential land. This may limit the capacity for less mobile fauna to colonise or persist on golf courses. Suburban golf courses might therefore be expected to have greater value for birds and animals with restricted area requirements than for less mobile fauna. An assessment of the relative ecological value of suburban golf courses that contrasts the level of utilisation by a range of vertebrates would present an opportunity to consider the way different animals respond to the landscape and thus, to assess the possible conservation value of small habitat remnants located in suburbia. 2.2 AIMS This chapter investigates the conservation of suburban golf courses in SEQ, Australia, by comparing vertebrate (bird, reptile, mammal and amphibian) assemblages on golf courses with those found in nearby eucalypt remnants and with bird assemblages in residential areas. Specifically this chapter aims to:
11 1) compare biodiversity (abundance, species richness and species diversity), 2) compare the abundance of individual species on each land type, 3) distinguish declining species from those that tolerate or thrive in urban areas, 4) compare species assemblages on each land type, 5) assess the extent to which golf courses act as a refuge for threatened species. 2.3 METHODS
2.3.1 SITE SELECTION Forty sites were surveyed for birds, reptiles, mammals and amphibians between August 2001 and March 2004. Sites included 20 golf courses, 10 residential areas and 10 eucalypt fragments, all located on the Gold Coast and Brisbane. All sites had been established for at least 20 years and occurred in flat, lowland areas with predominantly eucalypt vegetation (unburnt for at least 10 years). Most eucalypt fragments were comparable in size to golf courses (60ha), however some larger fragments (100-800ha) were included. Attempts were made to control for regional variation in biodiversity by selecting sites in regionally-paired groups, (one golf course, suburb and eucalypt fragment all within a 3km radius). This was however, not possible in all occasions, with three eucalypt fragments located further than 3km from their regional group. Ten additional unpaired golf courses were surveyed to increase sample size (Table 2.1). Table 2.1 Sites surveyed on the Gold Coast and Brisbane
BRISBANE Golf Eucalypt Forest Suburb Keperra Country Club Brisbane Forest Park Keperra/Ferny Hills Wynnum Golf Club Seven Hills Conservation Park Wynnum Redland Bay Golf Club Fragment: Lat: -27.64, Long: 153.28 Redland Bay The Pacific Golf Club Whites Hill Conservation Park Carindale California Creek Golf Club Venman State Forest Cornubia/Loganholme Oxley Golf Club Gailes Golf Club McLeods Golf Club St Lucia Golf Links Virginia Golf Club GOLD COAST Golf Eucalypt Forest Suburb The Grand Golf Club Nerang State Forest (West) Worongary Parkwood International Golf Club Fragment: Lat: -27.95, Long: 153.38 Parkwood Gold Coast-Burleigh Golf Club Clagiraba State Forest Miami/Burleigh Waters Helensvale Golf Club Coombabah Conservation Park Helensvale Southport Golf Club Nerang State Forest (East) Southport Gainsborough Greens Golf Club Surfers Paradise Golf Club Robina Woods Golf Club Tweed Heads-Coolangatta Golf Club Gold Coast Country Club
12
Brisbane
Residential Site Gold Coast
Eucalypt Site
Golf Course Site Urban Centres
Figure 2.1 Map of south-east Queensland, Australia showing regional distribution of study sites
13 2.3.2 SUB-SITE SELECTION On each golf course, eucalypt forest and suburb, terrestrial fauna surveys were conducted at ten sub-sites, randomly selected from numbered quadrants super-imposed over aerial photographs (Geoscape, 1997). On golf courses, all terrestrial sub-sites were located in rough and out-of-play areas. On each golf course and eucalypt forest, ten aquatic sub-sites were randomly selected from a list of all local waterbodies and surveyed for amphibians and wetland birds.
2.3.3 FAUNA SURVEYS Fauna surveys conducted in this study represent those that could be readily undertaken without extensive public consultation. All faunal groups (birds, reptiles, mammals and amphibians) were surveyed in eucalypt forests and on golf courses (with the permission of golf club management). Residential areas were however, surveyed for birds only.
BIRD SURVEYS All sites were surveyed for birds on six occasions, in Summer (2), Autumn (1), Winter (1) and Spring (2) between 2001-2003. Bird surveys were conducted on mornings without rainfall, within 3.5 hours of dawn. Each terrestrial and aquatic sub-site was surveyed for birds along a 100m x 30m strip transect, recording the number of birds seen 15m either side of the observer within a five-minute period. Surveys commenced after a five-minute settling period and did not count birds in flight.
REPTILE SURVEYS All golf courses and eucalypt forests were surveyed for reptiles on six occasions (in Spring and Summer) between 2001-2003. Each time, a one-hectare (ha) area in each sub-site was surveyed for 25 minutes using active search techniques. Five minutes was devoted to each of the following activities: 1) overturning rocks and logs 2) searching vertical substrates, 3) raking soil and leaf litter (in sunny conditions in early-mid morning and mid-late afternoon), 4) searching for basking reptiles in the heat of the day and 5) searching for nocturnal snakes. Each site was therefore surveyed for a total of 20 daylight hours and 5 hours at night. Geckos and turtles were not surveyed.
MAMMAL SURVEYS All golf courses and eucalypt forests were surveyed for mammals using a combination of spotlighting, Elliott trapping and opportunistic diurnal surveys. Twenty-minute spotlighting surveys were conducted along a 100m x 30m transect at each sub-site in Summer 2002 and Spring 2003. Each site was also surveyed for small ground mammals using baited Aluminium Elliott traps (33 x 10 x 9cm) with an effort of 90 trap nights
14 (30 traps: 3 per sub-site x 3 nights). Spotlighting and trapping surveys were conducted on clear nights with less than a half moon. Larger ground-dwelling mammals were recorded opportunistically, as encountered while conducting bird and reptile surveys. Bats (Microchiroptera and Megachiroptera) and mid-size nocturnal ground-dwelling mammals (eg. bandicoots) were not surveyed.
AMPHIBIAN SURVEYS All golf courses and eucalypt forests were surveyed for amphibians on three occasions (Summer 2002, Spring 2003 and Summer 2003) following rainfall events. At each aquatic sub-site, amphibians were surveyed in timed (15-minute) spotlighting searches of a 50m x 20m transect. All sites were surveyed within two weeks in any season to minimise variation in environmental conditions. Another three survey seasons were commenced, but abandoned due to variation in rainfall among sites. 2.4 DATA ANALYSIS
2.4.1 BIODIVERSITY ESTIMATES Site biodiversity indices including abundance, species richness, Simpson’s diversity index and Shannon-Weaver diversity index (Begon et al., 1986) were calculated for each vertebrate group. Biodiversity indices were compared among golf courses and eucalypt forests using t-tests (for reptiles, mammals and amphibians) and among golf courses, residential areas and eucalypt forests using one-way ANOVA’s (for birds). Tukey’s HSD (Zar, 1996) were used to identify significant differences.
2.4.2 SPECIES LAND PREFERENCES Site abundances of individual species were compared among land types, using the total cumulative site abundance (for bird, reptile and mammal species) and the maximum abundance recorded in any one survey (for amphibians). By comparing amphibian abundance under optimal conditions, the potential for variation due to environmental conditions (which can be problematic in short-term amphibian surveys) was reduced. Site abundances were compared using Kruskal-Wallis ANOVA (for birds) and Mann- Whitney U-tests (for reptiles, mammals and amphibians). Non-parametric methods were used due to a lack of homogeneity in the variance of some species.
2.4.3 ASSEMBLAGE COMPOSITION For each vertebrate group, species assemblages were compared among land types using semi-strong, hybrid multidimensional scaling ordination (SSH-MDS; Belbin, 1992) based on Bray Curtis dissimilarity measures (Bray and Curtis, 1957). These were calculated from data matrices that were range-standardised within sites to provide an
15 index of the relative abundance of each species at each site. Rare species (species with less than 5 individuals) were omitted from the analysis. A number of common species were dominant and therefore potentially obscured among-site variation in the relative abundance of less abundant species. Data were therefore fourth root transformed to create linearity and ensure all species contributed to the comparison among sites. In each ordination, stress was managed by adjusting the cut-point between metric and non-metric MDS. Principal axis correlation (PCC; Belbin, 1996) was used to identify species that contributed to the differences among sites. These species were plotted as vectors on the ordination axes and significance levels assessed using Monte- Carlo randomisation (MCAO; Belbin, 1996). Analysis of similarities (ANOSIM, Belbin, 1996) were used to test for differences in assemblage composition among land types. SIMPER analyses (Carr, 1996) were used to identify variables associated with significant differences among land types. Correlation analyses (Zar, 1996) were also used to test for associations between ordination axes and individual species abundance. SSH-MDS, ANOSIM, PCC and MCAO were conducted using the PATN statistical package (Belbin, 1996). SIMPER and ANOSIM (with Global R) analyses were conducted using the PRIMER package (Carr, 1996). 2.4.4 REFUGE VALUE To assess the capacity for golf courses to act as a refuge for threatened wildlife, it was first necessary to identify those species that are dependent on a source of refuge. All species were categorized into one of three groups based on categories proposed by Blair (1996): i.e. suburban ‘exploiting’, ‘tolerating’ and ‘avoiding’ species. Birds were assigned to categories by comparing their abundance in eucalypt forests and in suburban residential areas, using Mann-Whitney U-tests. ‘Exploiters’ were those significantly more abundant in residential areas than in eucalypt fragments. ‘Tolerators’ were those occurring in comparable abundance in both land types and ‘avoiders’ were those significantly less abundant in residential areas than in eucalypt remnants. Reptiles, mammals and amphibians could not be categorised in this way (given the lack of data for residential areas) and were therefore separated into suburban exploiters, tolerators and avoiders based on a mean ranking of the subjective opinion of six local wildlife biologists (i.e. Harry Hines from the Queensland Department of Natural Resources and Damian White, Jean-Marc Hero, Luke Shoo, Naomi Doak and Simon Hodgkison from Griffith University). Survey participants assigned each species a value from 1-5 where 1 = suburban exploiting, 2 = exploiting/tolerating, 3 = tolerating, 4 =
16 tolerating/avoiding and 5 = suburban avoiding (Appendix 1). Respondents only rated species with which they had previous field experience. Once likely suburban ‘avoiding’ species were identified, their combined site abundance was calculated from surveys conducted at each site and compared among land types using Mann-Whitney U-tests. 2.5 RESULTS
2.5.1 COMPARING BIODIVERSITY ESTIMATES Biodiversity - All Birds: 16,585 birds from 138 species were recorded in all land types. Birds were significantly more abundant on golf courses ( x = 452.3) than in eucalypt fragments ( x = 337.1, p = 0.006) and residential areas ( x = 361, p = 0.03; Fig. 2.2.a). Bird species richness was comparable on golf courses ( x = 49.4) and eucalypt fragments ( x = 47.2), but significantly lower in suburban areas ( x = 30.9; p < 0.0001; Fig. 2.2.b). Biodiversity - Wetland Birds: 2,534 individuals from 36 wetland bird species were found in all surveys. The vast majority of these were observed on golf courses. Golf courses supported a significantly higher abundance and species richness of aquatic birds than eucalypt fragments and suburban areas (Fig. 2.2.c,d). Biodiversity - Terrestrial Birds: 14,051 individuals from 102 terrestrial bird species were observed in all surveys. There was no significant difference in terrestrial bird abundance among land types (Fig. 2.2.e). Terrestrial bird species richness was however, significantly higher in eucalypt fragments ( x = 41.8), than in suburban areas ( x = 27.8, p = 0.003; Fig. 2.2.f). Biodiversity – Reptiles: 2,538 reptiles from 29 reptile species were recorded on golf courses and eucalypt forest fragments. There was no significant difference in the abundance of reptiles between golf courses ( x = 91.55) and eucalypt fragments ( x = 72.6, p = 0.3; Fig. 2.2.g). Eucalypt fragments had significantly higher levels of reptile species richness ( x euc = 9.8, x golf = 5.2, p < 0.0001; Fig. 2.2.h). Biodiversity – Amphibians: 8,777 individuals from 21 amphibian species were observed on golf courses and eucalypt forests. Amphibian abundance and species richness were significantly higher in eucalypt forests ( x Abund = 362.5, ( x Spp rich = 6.7) than on golf courses ( x Abund = 257.6, p = 0.05; x Spp rich = 4.5, p = 0.03; Fig. 2.2. i,j). Biodiversity – Mammals: 376 mammals from 17 species were observed on golf courses and in eucalypt forests. There was no significant difference in the abundance (p = 0.14) and species richness (p = 0.42) of mammals between golf courses and eucalypt fragments (Fig. 2.2.k,l).
17 2.5.2 LOCAL SPECIES ABUNDANCES Local Species Abundances – Birds: Thirty-eight terrestrial bird species displayed significant differences in abundance between land types. Twenty-three species were significantly more abundant in eucalypt forests than on golf courses and suburban areas. These included a range of small insectivorous forest birds (Table 2.2a-b). Only one species, (the galah), was significantly more abundant on golf courses than in other land types. Three common suburban species, (the Australian magpie, blue-faced honeyeater and house sparrow) were significantly more abundant in suburbia than in other land types. Eight other typically suburban birds (the noisy miner, figbird, crested pigeon, magpie-lark, willy wagtail, spotted turtle-dove, common myna and welcome swallow) were significantly more abundant on both golf courses and suburban areas than in eucalypt fragments (Table 2a-b). The dollarbird and superb blue wren were significantly more abundant on golf courses and eucalypt fragments than in suburbia. The noisy friarbird was significantly more abundant in residential areas and eucalypt forests than on golf courses. Twenty species were relatively common on all land types and are likely to be habitat generalists (Table 2.2a-b). Another 37 bird species were rare and therefore could not be analysed due to a lack of statistical power. Many of these are likely to be more abundant in forest fragments than on golf courses and residential areas. Local Species Abundances – Reptiles: Six reptile species displayed significant land preferences. The eastern water dragon, Physignathus lesueurii, was significantly more abundant on golf courses than in eucalypt fragments (U = 17.5, p = 0.001; Table 2.3). Species significantly more abundant in eucalypt fragments than on golf courses included the skinks Carlia vivax, (U = 1, p < 0.0001); Lampropholis amicula, (U = 18, p = 0.006); Ctenotus robustus, (U = 30, p = 0.03); and Ctenotus taeniolatus, (U = 28, p = 0.04) and the lace monitor, Varanus varius, (U = 16.5, p = 0.007). Species equally abundant on both land types included the skinks Lampropholis delicata, Cryptoblepharus virgatus, Eulamprus martini and Calyptotis scutirostrum, the bearded dragon Pogona barbata and the common tree snake Dendrelaphis punctulata (Table 2.3). Seventeen species were encountered in small numbers and therefore could not be assessed due to a lack of statistical power. Greater sampling effort, particularly for cryptic species and nocturnal snakes may have permitted a greater comparison of reptile habitat value among land types.
18 All Birds Aquatic Birds Terrestrial Birds
520 a) c) 420 e) 160
460 380
100
400 340
40 340 300 Abundance
280 -20 260 b) d) f) 55 14 42
10 45 36 6
35 30 2 Species Richness Richness Species
25 -2 24 Golf Forest Suburb Golf Forest Suburb Golf Forest Suburb
Reptiles Amphibians Mammals g) 130 i) 30 k)
450 110 20 350 90
10
Abundance 250 70
50 150 0
12 h) 9 j) 6 l)
8 10 5 7
8 6 4
5 6 3 4 Species Richness
4 3 2 Golf Forest Golf Forest Golf Forest FigureFigure 2.2 2.2 Mean Mean site site abundance abundance and and species species richness richness of ofall allbirds, birds, aquatic aquatic birds, birds, Terrestrialterrestrial birds, birds, reptiles, reptiles, mammals mammals and and amphibians amphibians in in residential residential areas (birdsareas onl (birdsy), golf only), courses golf andcourses eucal andyp teucalypt forests. Linesforests. = means,Lines = boxes means, = 1SE, whiskersboxes = = 1SE, 95% whiskers CI, shaded = 95% boxe CIs = significant differences. 19 Local Species Abundances – Amphibians: Five amphibian species displayed a significant difference in abundance between land types. Two species were significantly more abundant on golf courses than in eucalypt forests. These included the cane toad, Bufo marinus (U = 32, p = 0.003) and eastern sedge-frog, Litoria fallax (U = 20.5, p = 0.0004; Table 2.3). Three amphibian species were significantly more abundant in eucalypt fragments than on golf courses. These included the dainty green tree frog, Litoria gracilenta (U = 53, p = 0.01); common eastern froglet, Crinia signifera (U = 54.5, p = 0.004) and the copper-backed brood frog, Pseudophryne raveni (U = 10.5, p < 0.0001). Frogs that were found in comparable abundance on both land types included the striped marsh frog, Limnodynastes peronii; striped rocket frog, Litoria nasuta and the beeping froglet, Crinia parinsignifera. Fourteen rare species could not be assessed due to a lack of statistical power (Table 2.3). Local Species Abundances – Mammals: Only two mammal species displayed a significant difference in abundance between land types. The black rat Rattus rattus, was significantly more abundant on golf courses than in eucalypt fragments (U = 45, p = 0.02) and the yellow-footed antechinus Antechinus flavipes, was significantly more abundant in eucalypt forests (U = 49.5, p = 0.03) than on golf courses (Table 2.3). Seven mammal species were found in comparable abundance on both land types. These included the common brushtail possum Trichosurus vulpecula; common ringtail possum Pseudocheirus peregrinus; bush rat Rattus fuscipes, house mouse Mus musculus; koala Phascolarctos cinereus; red-necked wallaby Macropus rufogriseus and squirrel glider Petaurus norfolcensis (Table 2.3). Eight mammal species were found on fewer than six sites and were therefore not analysed due to a lack of statistical power. 2.5.3 ASSEMBLAGE COMPOSITION Assemblage Composition- All birds: Golf courses, suburban areas and eucalypt forests supported distinct bird species assemblages (Fig. 2.3). Analysis of similarity (ANOSIM), found a significant dissimilarity in bird assemblage among land types (ANOSIM, PATN p < 0.001; PRIMER Global R = 0.67, p = 0.001). Pair-wise comparisons indicated that eucalypt fragments support a significantly different bird assemblage than golf courses (ANOSIM: PRIMER, R = 0.76, p = 0.001) and suburban areas (ANOSIM: PRIMER, R = 0.97, p = 0.001). While there was some overlap in the bird assemblage on golf courses and suburban areas, the two were still significantly different (ANOSIM: PRIMER, R = 0.42, p = 0.001).
20 Table 2.2.a Total abundance of terrestrial bird species in each land type
LAND TYPE SUBURB GOLF FOREST N = 10 N = 20 N = 10 Species Common Name Total Sites Total Sites Total Sites P -Value RESPONSE GROUP
TERRESTRIAL BIRDS Manorina melanocephala Noisy Miner 524 10 1236 20 57 8 0.00006 GS Suburban Exploiting Streptopelia chinensis Spotted Turtle-Dove 393 10 123 17 23 5 0.00003 GS Suburban Exploiting Gymnorhina tibicen Australian Magpie 183 10 228 20 103 10 0.028 S Suburban Exploiting Ocyphaps lophotes Crested Pigeon 180 10 214 20 3 2 0.00002 GS Suburban Exploiting Grallina cyanoleuca Magpie-lark 180 10 185 20 23 7 0.00004 GS Suburban Exploiting Acridotheres tristis Common Myna 105 10 112 16 4 1 0.00009 GS Suburban Exploiting Sphecotheres viridis Figbird 93 9 213 18 6 4 0.005 GS Suburban Exploiting Passer domesticus House Sparrow 76 10 2 1 0 0 0.000001 S Suburban Exploiting Rhipidura leucophrys Willy Wagtail 47 10 162 19 2 2 0.00006 GS Suburban Exploiting Entomyzon cyanotis Blue-faced Honeyeater 32 8 4 3 0 0 0.0004 S Suburban Exploiting Sturnus vulgaris Common Starling 31 3 0 0 0 0 0.04 S Suburban Exploiting Columba livia Feral Pigeon 28 3 2 1 0 0 0.04 S Suburban Exploiting Hirundo neoxena Welcome Swallow 8 5 60 12 0 0 0.002 GS Suburban Exploiting Trichoglossus haematodus Rainbow Lorikeet 757 10 1383 20 522 10 NS Suburban Tolerating Trichoglossus chlorolepidotus Scaly-breasted Lorikeet 109 8 359 15 118 8 NS Suburban Tolerating Corvus orru Torresian Crow 183 10 342 19 217 9 NS Suburban Tolerating Cracticus nigrogularis Pied Butcherbird 119 10 283 19 100 10 NS Suburban Tolerating Cracticus torquatus Grey Butcherbird 79 10 204 18 76 9 NS Suburban Tolerating Lichmera indistincta Brown Honeyeater 62 7 167 12 39 6 NS Suburban Tolerating Cacatua roseicapilla Galah 28 6 116 15 6 3 0.03 G Suburban Tolerating Cacatua sanguinea Little Corella 9 5 106 10 3 1 NS Suburban Tolerating Zosterops lateralis Silvereye 20 7 87 8 75 8 NS Suburban Tolerating Platycercus adscitus Pale-headed Rosella 27 6 56 16 43 8 NS Suburban Tolerating Philemon corniculatus Noisy Friarbird 53 10 50 10 76 10 0.003 S F Suburban Tolerating Cacatua galerita Sulphur Crested Cockatoo 7 5 48 10 10 3 NS Suburban Tolerating Caracina novaehollandiae Black-faced Cuckoo-shrike 32 10 46 15 59 10 NS Suburban Tolerating Strepera graculina Pied Currawong 42 6 37 9 54 8 NS Suburban Tolerating Geopelia humeralis Bar-shouldered Dove 2 2 26 9 24 4 NS Suburban Tolerating Eudynamis scolopacea Common Koel 5 4 24 12 3 2 NS Suburban Tolerating Anthochaera lunulata Little Wattlebird 8 3 9 3 3 3 NS Suburban Tolerating Milvus sphenurus Whistling Kite 2 1 9 4 5 5 NS Suburban Tolerating Malurus assimilis Variegated Wren 3 1 53 9 45 5 NS Suburban Tolerating Psophodes olivaceus Eastern Whipbird 3 2 50 9 43 5 NS Suburban Tolerating Myzomela sanguinolenta Scarlet Honeyeater 7 4 47 6 210 9 0.0003 F Suburban Avoiding Colluricincla harmonica Grey Shrike-thrush 7 3 62 10 138 10 0.00004 F Suburban Avoiding Pardalotus striatus Striated Pardalote 48 9 88 15 138 10 0.0008 F Suburban Avoiding Rhipidura fuliginosa Grey Fantail 4 2 39 3 110 9 0.00003 F Suburban Avoiding Melithreptus albogularis White-throated Honeyeater 4 2 23 9 87 10 0.00002 F Suburban Avoiding Gerygone olivacea White-throated Gerygone 3 2 26 6 64 8 0.001 F Suburban Avoiding Cormobates leucophaea White-throated Treecreeper 1 1 10 6 61 10 0.00002 F Suburban Avoiding Merops ornatus Rainbow Bee-eater 2 2 18 4 58 8 0.001 F Suburban Avoiding Oriolus sagittatus Olive-backed Oriole 16 6 67 15 57 10 0.008 F Suburban Avoiding Pachycephala rufiventris Rufous Whistler 0 0 37 6 57 8 0.0005 F Suburban Avoiding Acanthiza reguloides Buff-rumped Thornbill 0 0 0 0 46 3 0.04 F Suburban Avoiding Malurus melanocephalus Red-backed Wren 0 0 18 4 45 7 0.001 F Suburban Avoiding Todiramphus sancta Sacred Kingfisher 4 3 57 14 44 7 0.05 F Suburban Avoiding Malurus cyaneus Superb Blue Wren 0 0 110 11 44 5 0.03 F Suburban Avoiding Lichenostomus chrysops Yellow-faced Honeyeater 2 1 17 4 42 5 0.03 F Suburban Avoiding Dacelo noveguineae Laughing Kookaburra 10 4 62 14 39 10 0.03 F Suburban Avoiding Pachycephala pectoralis Golden Whistler 0 0 14 3 38 7 0.0002 F Suburban Avoiding Poephila bichenovii Double-barred Finch 0 0 14 4 37 4 0.05 F Suburban Avoiding Acanthiza pusilla Brown Thornbill 0 0 32 7 34 8 0.002 F Suburban Avoiding Dicrurus bracteatus Spangled Drongo 4 2 19 7 31 8 0.006 F Suburban Avoiding Coracina tenuirostris Cicadabird 0 0 10 5 25 7 0.001 F Suburban Avoiding Rhipidura rufifrons Rufous Fantail 0 0 6 3 22 4 0.01 F Suburban Avoiding Eurystomus orientalis Dollarbird 0 0 50 13 16 7 0.003 F Suburban Avoiding Geopelia striata Peaceful Dove 0 0 5 3 16 5 0.03 F Suburban Avoiding Sericornis frontalis White-browed Scrubwren 0 0 7 3 15 5 0.01 F Suburban Avoiding Cuculus flabelliformis Fan-tailed Cuckoo 0 0 3 2 12 5 0.003 F Suburban Avoiding Myiagra rubecula Leaden Flycatcher 0 0 3 1 12 4 0.01 F Suburban Avoiding Cinclosoma punctatum Spotted Quail-Thrush 0 0 0 0 11 3 0.04 F Suburban Avoiding Smicrornis brevirostris Weebill 0 0 5 1 11 3 0.04 F Suburban Avoiding Myiagra inquieta Restless Flycatcher 0 0 5 3 8 4 0.04 F Suburban Avoiding Eopsaltria australis Eastern Yellow Robin 0 0 3 1 5 3 0.04 F Suburban Avoiding Meliphaga lewinii Lewin's Honeyeater 0 0 45 6 20 4 NS Insufficient Power Todiramphus macleayii Forest Kingfisher 0 0 11 5 7 4 NS Insufficient Power
P – Values represent results from Kruskal-Wallis comparison of abundance among land types. Initials represent land types in which each species was most significantly abundant. G = Golf courses, S = Suburbia, F = Eucalypt forest. Classifications based on Mann- Whitney U test comparing abundance in residential areas and eucalypt forests. Suburban Exploiting (significantly more abundant in suburbia), Suburban Tolerating (found in comparable abundance in both land types), Suburban Avoiding (significantly more abundant in eucalypt forest).
21 Table 2.2.b Total abundance of terrestrial and aquatic birds in each land type
LAND TYPE SUBURB GOLF FOREST N = 10 N = 20 N = 10 Species Common Name Total Sites Total Sites Total Sites P-Value
TERRESTRIAL BIRDS Neochmia temporalis Red-browed Firetail 2 1 32 1 5 2 No Power Alectura lathami Australian Brush-turkey 0 0 22 3 0 0 No Power Lonchura castaneothorax Chestnut-breasted Mannikin 0 0 20 2 0 0 No Power Centropus phasianinus Pheasant Coucal 0 0 15 7 3 3 No Power Philemon citreogularis Little Friarbird 6 3 14 4 0 0 No Power Scythrops novaehollandiae Channel-billed Cuckoo 0 0 13 6 6 4 No Power Pomatostomus temporalis Grey Crowned Babbler 0 0 12 1 0 0 No Power Myiagra cyanoleuca Satin Flycatcher 0 0 10 2 2 2 No Power Alisterus scapularis Australian King Parrot 1 1 9 1 5 3 No Power Todiramphus chloris Collared Kingfisher 0 0 9 4 0 0 No Power Falco cenchroides Nankeen Kestrel 0 0 9 7 1 1 No Power Sericornis magnirostris Large-billed Scrubwren 0 0 8 3 4 2 No Power Gerygone laevigaster Mangrove Gerygone 0 0 7 2 3 2 No Power Chrysoccyx basalis Horsfield's Bronze Cuckoo 0 0 6 4 6 2 No Power Platycercus eximius Eastern Rosella 0 0 4 3 0 0 No Power Podargus strigoides Tawny Frogmouth 0 0 4 4 1 1 No Power Ceyx azura Azure Kingfisher 0 0 3 2 0 0 No Power Cuculus variolosus Brush Cuckoo 0 0 3 1 3 2 No Power Accipiter cirrhocephalus Collared Sparrowhawk 0 0 3 1 1 1 No Power Cisticola exilis Golden-headed Cisticola 0 0 3 1 2 1 No Power Dicaecum hirundinaceum Mistletoebird 0 0 3 2 1 1 No Power Pardalotus punctatus Spotted Pardalote 0 0 3 2 0 0 No Power Hirundo nigricans Tree Martin 0 0 3 1 0 0 No Power Macropygia amboinensis Brown Cuckoo-dove 0 0 2 1 2 2 No Power Climacteris picumnus Brown Treecreeper 0 0 2 1 3 2 No Power Ninox novaeseelandiae Southern Boobook 0 0 2 1 0 0 No Power Lalage leucomela Varied Triller 0 0 2 1 0 0 No Power Falco berigora Brown Falcon 0 0 1 1 0 0 No Power Aviceda subcristata Pacific Baza 0 0 1 1 0 0 No Power Manorina melanophrys Bell Miner 2 1 0 0 3 1 No Power Catyptorhynchus lathami Glossy Black Cockatoo 0 0 0 0 3 1 No Power Turnix varia Painted Button Quail 0 0 0 0 4 2 No Power Falco peregrinus Peregrine Falcon 0 0 0 0 1 1 No Power Ninox strenua Powerful Owl 0 0 0 0 1 1 No Power Plectorhyncha lanceolata Striped Honeyeater 0 0 0 0 2 1 No Power AQUATIC BIRDS Gallinula tenebrosa Dusky Moorhen 0 0 407 14 0 0 0.0002 G Chenonetta jubata Australian Wood Duck 0 0 371 18 1 1 0.0001 G Fulica atra Eurasian Coot 0 0 286 8 0 0 0.003 G Porphyrio porphyrio Purple Swamphen 0 0 215 11 3 1 0.01 G Anas superciliosa Pacific Black Duck 2 3 204 17 1 1 0.0002 G Threskiornis aethiopica Sacred Ibis 45 0 202 10 1 1 0.01 G Aythya australis Hardhead 0 0 160 14 0 0 0.0008 G Threskiornis spinicollis Straw-necked Ibis 0 0 145 10 0 0 0.01 G Vallenus miles Masked Lapwing 6 2 123 15 0 0 0.0003 G Phalacrocorax sulcirostris Little Black Cormorant 0 0 88 11 0 0 0.003 G Anas castanea Chestnut Teal 0 0 47 6 0 0 0.04 G Tachybaptus novaehollandiae Australasian Grebe 0 0 45 13 0 0 0.0008 G Ardea novaehollandiae White-faced Heron 1 1 24 13 0 0 0.0001 G Anas platyrhynchos Mallard 0 0 22 6 0 0 0.04 G Platalea regia Royal Spoonbill 0 0 20 9 0 0 0.01 G Phalacrocorax melanoleucos Little Pied Cormorant 0 0 19 8 0 0 0.003 G Anhinga melanogaster Australian Darter 0 0 10 4 0 0 No Power Anas gracilis Australian Grey Teal 0 0 9 3 0 0 No Power Gallinago hardwickii Japanese Snipe 0 0 9 2 0 0 No Power Nycticorax caledonicus Nankeen Night Heron 0 0 9 7 0 0 No Power Cygnus olor Black Swan 0 0 8 3 0 0 No Power Ardea intermedia Intermediate Egret 0 0 8 5 0 0 No Power Ardea garzetta Little Egret 0 0 8 5 0 0 No Power Ardea alba Great Egret 1 2 7 6 0 0 No Power Himantopus himantopus Black-winged Stilt 0 0 4 1 0 0 No Power Jacana gallinacea Comb-crested Jacana 0 0 4 3 0 0 No Power Dendrocygna eytoni Plumed Whistling Duck 0 0 4 1 0 0 No Power Ardea ibis Cattle Egret 0 0 3 2 0 0 No Power Rallus philippensis Buff-banded Rail 0 0 2 1 0 0 No Power Burhinus grallarius Bush Thick-knee 0 0 2 2 0 0 No Power Ardeola striatus Mangrove Bittern 0 0 2 1 0 0 No Power Charadrius melanops Back-fronted Dotterel 0 0 1 1 0 0 No Power Plegadis falcinellus Glossy Ibis 0 0 1 1 0 0 No Power Dendrocygna arcuata Wandering Whistling Duck 0 0 1 1 0 0 No Power Pelecanus conspicillatus Australian Pelican 1 1 0 0 0 0 No Power Ardea pacifica Pacific Heron 0 0 0 0 1 1 No Power P – Values = results from Kruskal-Wallis comparison of abundance among land types. Initials = land types where most significantly abundant. G = Golf, S = Suburbia, F = Forest. Classifications based on Mann-Whitney U test comparing abundance in residential areas and eucalypt forests. Suburban Exploiting (significantly more abundant in suburbia), Suburban Tolerating (comparable abundance in both land types), Suburban Avoiding (significantly more abundant in eucalypt forest).
22 Table 2.3. Total abundance of reptile, amphibian and mammal species in each land type
LAND TYPE GOLF FOREST N = 20 N = 10 Species Common Name Total Sites Total Sites P-Value Response Group
REPTILES Lampropholis delicata Wall skink 345 20 153 10 NS Suburban Exploiting Cryptoblepharus virgatus Grass skink 898 20 347 10 NS Suburban Exploiting Physignathus lesueurii Eastern water dragon 413 16 11 3 0.001 G Suburban Tolerating Ctenotus robustus Robust skink 1 1 4 4 0.03 F Suburban Tolerating Anomalopus verreauxii Verreaux's skink 1 1 4 4 NS Suburban Tolerating Eulamprus quoyii Eastern water skink 21 4 0 0 NS Suburban Tolerating Pogona barbata Bearded dragon 42 13 13 5 NS Suburban Tolerating Calyptotis scutirostrum Scaly-Snouted skink 32 10 15 6 NS Suburban Tolerating Tiliqua scincoides Blue-tongued skink 1 1 2 2 No Power Suburban Tolerating Carlia vivax Lively skink 8 4 72 10 0.00001 F Suburban Avoiding Lampropholis amicula Secretive skink 13 6 23 7 0.006 F Suburban Avoiding Varanus varius Lace monitor 8 4 29 8 0.007 F Suburban Avoiding Ctenotus taeniolatus Copper Tailed skink 6 4 10 5 0.04 F Suburban Avoiding Carlia schmeltzii Robust Rainbow skink 3 1 3 2 NS Suburban Avoiding Eulamprus martini Martin's skink 4 4 11 5 NS Suburban Avoiding Ramphotyphlops nigrescens Blind snake 0 0 7 2 NS Suburban Avoiding Lialis burtonis Burton's legless lizard 2 2 5 3 NS Suburban Avoiding Dendrelaphis punctulata Common tree snake 9 6 3 3 NS Suburban Avoiding Rhinoplocephalus nigrescens Eastern Small-eyed snake 0 0 2 2 No Power Suburban Avoiding Morelia spilota Carpet python 3 3 0 0 No Power Suburban Avoiding Amphibolurus muricatus Jacky lizard 0 0 2 1 No Power Suburban Avoiding Coeranoscincus reticulatus Three-toed snake-tooth skink 0 0 1 1 No Power Suburban Avoiding Ctenotus arcanus No Common Name 0 0 3 1 No Power Suburban Avoiding Demansia psammophis Yellow-faced Whip snake 0 0 1 1 No Power Suburban Avoiding Diporiphora australis Eastern Two-lined dragon 0 0 1 1 No Power Suburban Avoiding Eulamprus tenuis Bar-sided skink 0 0 3 1 No Power Suburban Avoiding Pseudechis porphyriacus Red-bellied black snake 0 0 1 1 No Power Suburban Avoiding Ophioscincus truncatus Short-limbed Snake skink 1 1 0 0 No Power Suburban Avoiding Tropidonophis mairii Keelback snake 1 1 0 0 No Power Suburban Avoiding AMPHIBIANS Bufo marinus Cane Toad 1747 20 445 10 0.001 G Suburban Exploiting Limnodynastes peronii Striped Marsh Frog 489 18 672 7 NS Suburban Exploiting Litoria fallax Eastern Sedge Frog 2348 18 126 6 0.01 G Suburban Tolerating Litoria gracilenta Dainty Green Tree Frog 167 3 386 6 0.01 F Suburban Tolerating Crinia parinsignifera Beeping Froglet 40 3 143 5 NS Suburban Tolerating Litoria caerulea Green Tree Frog 4 2 0 0 No Power Suburban Tolerating Pseudophryne raveni Copper-backed Brood Frog 1 1 1515 9 0.00001 F Suburban Avoiding Litoria nasuta Rocket Frog 225 3 28 5 NS Suburban Avoiding Litoria latopalmata Broad Palmed frog 26 2 7 2 NS Suburban Avoiding Adelotus brevis Tusked Frog 27 5 12 2 NS Suburban Avoiding Crinia signifera Common Eastern Froglet 3 1 96 5 NS Suburban Avoiding Litoria peronii Peron's Tree Frog 45 6 3 1 NS Suburban Avoiding Pseudophryne major Great Brown Brood Frog 0 0 118 2 No Power Suburban Avoiding Crinia tinnula Wallum Froglet 11 1 57 1 No Power Suburban Avoiding Mixophyes fasciolatus Great Barred Frog 0 0 10 3 No Power Suburban Avoiding Limnodynastes dumerilii Eastern Banjo Frog 2 1 3 1 No Power Suburban Avoiding Limnodynastes salmini Salmon-striped Frog 0 0 2 1 No Power Suburban Avoiding Uperoleia fusca Dusky Toadlet 1 1 2 1 No Power Suburban Avoiding Limnodynastes ornatus Ornate Burrowing Frog 5 1 0 0 No Power Suburban Avoiding Litoria dentata Bleating Tree Frog 10 2 0 0 No Power Suburban Avoiding Litoria tyleri Tyler's Tree Frog 1 1 0 0 No Power Suburban Avoiding MAMMALS Rattus rattus Black Rat 31 11 0 0 0.02 G Suburban Exploiting Mus musculus House Mouse 20 10 2 1 NS Suburban Exploiting Trichosurus vulpecula Common Brushtail Possum 44 13 32 8 NS Suburban Tolerating Pseudocheirus peregrinus Common Ringtail Posum 18 6 3 2 NS Suburban Tolerating Macropus rufogriseus Red-necked Wallaby 12 4 14 5 NS Suburban Avoiding Rattus fuscipes Bush Rat 8 4 5 3 NS Suburban Avoiding Petaurus norfolcensis Squirrel Glider 2 2 4 4 NS Suburban Avoiding Antechinus flavipes Yellow-footed Antechinus 3 2 10 6 0.03 F Suburban Avoiding Macropus giganteus Eastern Grey Kangaroo 116 3 4 1 No Power Suburban Avoiding Phascolarctos cinereus Koala 7 2 8 4 No Power Suburban Avoiding Rattus tunneyi Pale Field Rat 1 1 2 1 No Power Suburban Avoiding Petauroides volans Greater Glider 0 0 4 2 No Power Suburban Avoiding Rattus lutreolus Swamp Rat 4 3 1 1 No Power Suburban Avoiding Wallabia bicolor Swamp Wallaby 2 2 0 0 No Power Suburban Avoiding Lepus capensis Brown Hare 3 3 0 0 No Power Suburban Avoiding Canis lupus familiaris Wild Dog 0 0 1 1 No Power Suburban Avoiding Vulpes vulpes European Fox 2 2 2 2 No Power Suburban Avoiding
P – Values represent results from Mann-Whitney U test comparison of abundance among land types. Initials represent land types in which each species was significantly most abundant. G = Golf courses, S = Suburbia, F = Eucalypt forest. Classifications represents predicted status, based on the opinion of six local SEQ wildlife biologists.
23 Suburban areas displayed less within variation than golf courses and eucalypt forests, indicating a higher level of homogenisation. Suburban bird assemblages shared 71% compositional similarity, compared to 63% within eucalypt forests and 57% within golf courses (SIMPER). The separation of sites was driven by species including the dusky moorhen (r = 0.80) and Pacific black duck (r = 0.87), which were relatively abundant on golf courses, and the white-throated honeyeater (r = 0.88), grey shrike- thrush (r = 0.89), striated pardalote (r = 0.79), scarlet honeyeater (r = 0.85) and noisy friarbird (r = 0.75), which were more abundant on eucalypt fragments (PCC Vectors). Assemblage Composition - Terrestrial Birds: Aquatic birds accounted for much of the variation in bird species assemblages. Thus when aquatic species were removed there was greater overlap among land types (Fig. 2.3). Nevertheless, there were still significant differences in terrestrial bird assemblages among land types (ANOSIM, PATN: p < 0.0001, PRIMER Global R = 0.53, p = 0.001). Assemblages on eucalypt fragments were significantly different from those found in suburban areas (ANOSIM R = 0.96, p = 0.001) and golf courses (ANOSIM R = 0.74, p = 0.001). While the terrestrial bird assemblage on golf courses were also significantly different to those found in suburban areas (ANOSIM R = 0.14, p = 0.05), the low R-value indicates they were only marginally separable. Within land type variation was greatest for eucalypt fragments, intermediate for golf courses and most restricted for suburban areas, indicating increasing homogenisation as land types become more urbanised. This was reflected in the SIMPER results. Suburban bird assemblages shared 72% compositional similarity, compared with 59% within golf courses and eucalypt forests. The separation of sites was driven by species including the white-throated honeyeater (r = 0.89), cicadabird (r = 0.83) and scarlet honeyeater (r = 0.87) which were more abundant in eucalypt forests, the noisy miner (r = 0.87) which was more abundant in suburban areas and the crested pigeon (r = 0.79) and magpie-lark (r = 0.83) which were more abundant on golf courses (PCC; Fig. 2.3). Assemblage Composition – Reptiles: Reptile assemblages on golf courses and eucalypt forests displayed considerable variation and overlap (Fig. 2.3). Reptile assemblages on golf courses were nevertheless, significantly different from those found in eucalypt fragments (ANOSIM, PATN p < 0.0001; PRIMER Global R = 0.46, p = 0.001). Variation within land types was comparable for both eucalypt fragments (SIMPER similarity = 65%) and golf courses (SIMPER similarity = 62%). Species
24 driving the separation of sites included the water dragon, Physignathus lesueurii, (PCC r = 0.94) and skinks including Carlia vivax, (PCC r = 0.86); Lampropholis amicula, (PCC r = 0.76) and Calyptotis scutirostrum (PCC r = 0.80; Fig. 2.3). Reptiles correlated with ordination Axis 1 included the lace monitor Varanus varius, (r = -0.68, p < 0.0001), lively skink, Carlia vivax, (r = -0.83, p < 0.0001) and the eastern water dragon Physignathus lesueurii, (r = 0.90, p < 0.0001). Species correlated with ordination Axis 2 included the skinks Calyptotis scutirostrum, (r = 0.79, p < 0.0001) and Ctenotus robustus, (r = -0.39, p = 0.03). Assemblage Composition – Amphibians: Golf courses supported amphibian assemblages that were significantly different to those found in eucalypt fragments (ANOSIM - PATN p < 0.001, PRIMER Global R = 0.76, p = 0.001; Fig. 2.3). These differences were largely defined by ordination Axis 1, which separated sites (i.e. golf courses) with a relative abundance of Bufo marinus (r = -0.68, p < 0.0001) and Litoria fallax (r = -0.52, p = 0.003), from sites (i.e. eucalypt fragments) that had greater proportions of Pseudophryne raveni (r = 0.87, p < 0.0001), Crinia signifera (r = 0.61, p = 0.0003), Litoria gracilenta (r = 0.53, p = 0.002) and Crinia parinsignifera (r = 0.52, p = 0.003; Fig. 2.3). In contrast, ordination Axis 2 largely defined within-land-type variation. Four species were significantly correlated with ordination Axis 2: (Litoria gracilenta, r = 0.51, p = 0.004; Litoria fallax, r = -0.78, p < 0.0001; Crinia tinnula, r = -0.41, p = 0.02 and Litoria peronii, r = -0.38, p = 0.04). Principal axis correlations (PCC) supported these results, suggesting that the ordination was driven by species including Pseudophryne raveni (r = 0.89), Crinia parinsignifera (r = 0.73), Litoria gracilenta (r = 0.68), Bufo marinus (r = 0.69) and Litoria fallax (r = 0.88). Amphibian assemblages on golf courses were considerably more predictable than those in eucalypt forests. Three species (Bufo marinus, Litoria fallax and Limnodynastes peronii) contributed 96.9% of the similarity in amphibian assemblage on golf courses. In contrast, six species including (Pseudophryne raveni, Bufo marinus, Limnodynastes peronii, Litoria gracilenta, Litoria fallax and Crinia parinsignifera) explained 90.7% of the similarity in amphibian species assemblages in eucalypt forests. There was very little overlap in the amphibian assemblage on golf courses and eucalypt forests. Only two of the twenty golf courses surveyed (G4 and G19) supported an amphibian assemblage comparable to those found in eucalypt forests (Fig. 2.3). Assemblage Composition – Mammals: Mammal assemblages varied substantially among sites. Three dimensions were required to represent that variation while
25 maintaining acceptable stress levels. Despite substantial overlap in mammal assemblages on golf courses and eucalypt forests (Fig. 2.3) assemblages were significantly different (ANOSIM – PRIMER, Global R = 0.24, p = 0.002). Differences between the two land types were evident in the Axis 1–Axis 2 observation plane. Golf courses were distinguished from eucalypt forests by a relatively high abundance of the black rat, Rattus rattus (PCC, r = 0.87) and house mouse, Mus musculus (PCC, r = 0.80). Other species highly correlated with the ordination matrix (PCC) included the brushtail possum, Trichosurus vulpecula (r = 0.87); eastern grey kangaroo, Macropus giganteus (r = 0.72); yellow-footed antechinus, Antechinus flavipes (r = 0.62); bush rat, Rattus fuscipes (r = 0.61) and fox, Vulpes vulpes (r = 0.61). 2.5.4 SUBURBAN ‘EXPLOITING, AVOIDING AND TOLERATING’ SPECIES Mann-Whitney U tests comparing bird species abundances in suburban areas and eucalypt forests identified 13 suburban-exploiting, 20 suburban-tolerating and 31 suburban-avoiding species. Another 37 bird species were rare in both land types and could therefore not be assessed due to a lack of statistical power (Table 2.2b). Reptiles, mammals and amphibians were categorized into suburban exploiting, tolerating and avoiding species based on the expert opinion of six local wildlife biologists. Reptiles were separated into 2 suburban exploiting, 7 tolerating and 20 avoiding species. Frogs were divided into 2 suburban exploiting, 4 tolerating and 15 avoiding species. Mammal groups consisted of 2 exploiting, 2 tolerating and 10 avoiding species (Appendix 1).
2.5.5 REFUGE VALUE As expected, eucalypt forests were a better source of refuge for regionally threatened birds than golf courses and suburban areas. Golf courses varied substantially in the extent to which they provide habitat for regionally threatened wildlife. Golf courses were generally a better refuge for birds and mammals than for reptiles and amphibians (Fig. 2.4). Threatened mammals were found in densities, comparable to those in eucalypt forest on seven golf courses (G1,2,5,17,18,19,20). Four golf courses (G3,13,18,19) were a substantial refuge for regionally threatened birds. Only two courses (G18,19) were a refuge for threatened amphibians and 3-4 courses (G1,5,17,19) supported threatened reptiles in densities comparable to those found in low quality eucalypt remnants. Most golf courses had negligible refuge value, supporting primarily common urban-adapted species.
26 All Birds Terrestrial Birds Stress = 0.1841 Stress = 0.1968
Pacific Black Duck Noisy Miner Dusky Moorhen Crested Pigeon White- Crested throated Pigeon Magpie Honeyeater White- throated Lark Honeyeater Grey Shrike- thrush Cicadabird Striated Noisy Friarbird Scarlet Pardalote Honeyeater Striated
Axis 2 Pardalote Axis 2 Axis 1 Axis 1 Reptiles Amphibians Stress = 0.2028 Stress = 0.1806
Calyptotis scutirostrum Litoria gracilenta
Pseudophryne Bufo raveni marinus
Carlia vivax
Physignathus Lampropholis Crinia lesueurii Litoria fallax amicula parinsignifera Axis 2 Axis 2 Axis 1 Axis 1 Mammals Mammals Stress = 0.1930 Stress = 0.1930 Common Brushtail House Mouse Possum Black Rat Common Brushtail Black Rat Red-necked Eastern Grey Possum Wallaby Kangaroo
Yellow-footed Red-necked Antechinus Bush House Mouse Wallaby Eastern Grey Rat Kangaroo Yellow-footed Bush Rat Antechinus Axis 2 Axis 3 Axis 1 Axis 1 Figure 2.3. SSH-MDS ordination plots showing the relative similarity in species composition among sites for each vertebrate group. (Red circles = eucalypt forest, blue triangles = golf courses, green squares = residential areas, ellipses = 90% CI, arrows = significant PCC vectors). 27 100 Birds 80
60
40
20 Mean Abundance
180 Amphibians
140
100 No Data
60
Mean Abundance 20
Reptiles 8
6 No Data
4
Mean Abundance 2
2 Mammals
No Data 1
Log Total Abundance 0
F5 F7 F9 F3 F6 F8 F1 F2 F4 S1 S7 S8 S4 S3 S2 S9 S5 S6 G3 G1 G2 G4 G8 G5 G7 G9 G6 F10 S10 G18 G19 G13 G16 G14 G17 G12 G20 G11 G15 G10
Forests Golf Courses Suburbs
Figure 2.4 The refuge value of individual sites indicated by the mean site abundance of threatened birds, reptiles, mammals and amphibians. Columns = means, whiskers = 95% CI. (For mammals columns = 28 total from all mammal surveys). 2.6 DISCUSSION Urban planners require information on the ecological value of all urban land types in order to provide an ecological basis for urban zoning decisions; to ensure formal conservation reserves are bordered by appropriate land types and to identify opportunities for off-reserve conservation (Marzluff and Ewing, 2001). The results of this study indicate that urban planners would currently have little ability to predict the ecological value of golf courses, based simply on their land-use. Golf courses displayed extreme variation in biodiversity. Similar variation has been observed in bird assemblages on golf courses in the USA (Hostetler and Knowles-Yanez, 2003; Jones et al., 2005; LeClerc and Cristol, 2005; Merola-Zwartjes and DeLong, 2005; Porter et al., 2005). Importantly, the results of this study indicate that suburban golf courses can provide habitat for a range of regionally threatened vertebrates and therefore have localised conservation value. However, while golf courses evidently have the capacity to act as a refuge for threatened wildlife, most only support common urban-adapted species and therefore fail to realise that potential. Wildlife assemblages on golf courses and in residential areas were generally more homogenised than those in eucalypt fragments, supporting higher abundances, but lower species richness for most vertebrate groups. Community-level homogenisation is typically observed in degraded landscapes and generally occurs when a limited number of persistent species (able to tolerate the modified conditions), occur in high local densities (Blair, 1996; McKinney and Lockwood, 1999; Blair, 2001). Bird assemblages on golf courses were less homogenised than those in suburbia and while reptile, mammal and amphibian assemblages were not directly compared between golf courses and residential areas, the utilisation of golf courses by species rarely observed in suburban areas of SEQ (eg. koalas, macropods, gliders, lace monitors, forest-dwelling skinks and ephemeral pond-breeding amphibians), suggests some golf courses are a superior source of habitat to residential areas. Like urban waste-water settling ponds (Greenway and Simpson, 1996) and storm-water mitigation areas (McGuckin and Brown, 1995), pond networks retained on suburban golf courses in SEQ appear to have value as a source of urban wetland habitat for aquatic birds. Golf courses in Florida, USA have been found to perform a similar role (White and Main, 2005). The conservation value of aquatic habitats on golf courses in SEQ could however not be determined by comparison with eucalypt forests, since these are not a typical source of waterbird habitat. Wetland bird assemblages on golf
29 courses were dominated by a small suite of relatively urban-adapted species (eg. the Eurasian coot, dusky moorhen, Pacific black duck, Australian wood duck and purple swamphen) and may therefore be more homogenised than aquatic bird assemblages in natural wetlands. Nevertheless, regionally uncommon species including the comb- crested jacana, wandering whistling duck, plumed whistling duck, buff-banded rail, black swan, and Japanese snipe were observed (some nesting) on a number of golf courses, suggesting golf courses do have some conservation value for aquatic birds. 2.6.1 IMPLICATIONS FOR NEW GOLF COURSE CONSTRUCTION The results support previous studies that have stressed the need to use appropriate selection criteria when assessing ecological value (Margules and Usher, 1981; Margules and Nicholls, 1987) and consider the consequences of landscape change at more than one scale (Savard et al., 2000; Hazell et al., 2001). Some golf courses supported higher bird species richness (terrestrial and aquatic) than eucalypt forests. The replacement of eucalypt forest with a new golf course could therefore in some circumstances, lead to a local increase in bird species richness. However, since most birds occurring on golf courses are regionally common and the majority of birds found in eucalypt fragments are regionally declining, it cannot be suggested that golf courses are ecologically superior to eucalypt remnants or that the replacement of forest by golf courses would result in a positive ecological outcome. While the construction of a new golf course may result in increased biodiversity at a local level, it would inevitably contribute to reduced biodiversity on a regional scale. Urban land-use decisions must therefore look beyond local and immediate changes in biodiversity and consider the consequences at other spatial and temporal scales. Ultimately, the ecological consequences of any golf course construction will depend on the type of land that is lost to development. Any development that results in substantial loss of eucalypt forest will lead to the local decline of many regionally threatened species.
2.6.2 REFUGE-VALUE OF GOLF COURSES Golf courses were however, not without ecological value. A small number of golf courses retained threatened suburban-avoiding wildlife. Threatened species found on golf courses included a range of small insectivorous birds, three macropod species, three species of native rats, the koala, squirrel glider, yellow-footed antechinus and fourteen forest-dependent reptile and amphibian species. The results confirm suggestions (Maffei, 1978; Tatnall, 1991; Tietge, 1992; Pearce, 1993; Terman, 1997; Dawson, 2000; Cristol and Rodewald, 2005; Merola-Zwartjes and DeLong, 2005) that
30 golf courses have the potential to act as a refuge for threatened wildlife. Golf courses that were a refuge for one vertebrate group were however, not necessarily of value to others. This was not unexpected, since animals are known to respond to different landscape features, depending on their morphology, physiology, home range and behaviour (Addicott et al., 1987; Wiens, 1989). Interestingly, golf courses were a better refuge for threatened birds and mammals than for threatened reptiles and amphibians. The relative failure of golf courses to support threatened herpetofauna could be attributed to a number of factors. Many amphibians have limited mobility (Harris, 1975; Beshkov and Jameson, 1980; Sinsch, 1990; Ficetola and De Bernardi, 2004), and can therefore be particularly susceptible to the isolating effects of habitat fragmentation, given their inability to overcome inter-remnant gaps (Laan and Verboom, 1990; Sjögren, 1991; Blaustein et al., 1994; Marsh and Pearman, 1997). Many golf courses in SEQ are isolated by suburbia. This may inhibit the capacity for some amphibians and reptiles to colonize, recolonise or persist on isolated suburban golf courses. While large, ground-dwelling mammals would also be susceptible to entrapment on isolated golf courses, they may have a greater capacity for short-term persistence, given their relative longevity (and thus reduced rate of population turnover). Alternatively, the relative absence of threatened herpetofauna from golf courses could be attributed to a difference in the extent to which their habitats have been compromised or to increased exposure to local threats. Exposure to atrazine and other chemical herbicides can reduce amphibian reproductive success (Diana et al., 2000; Hayes et al., 2002) and may suppress native frog populations on golf courses (where herbicides are still widely used). Reptiles and amphibians may also suffer higher predation pressures, given the relatively high density of avian predators on golf courses and the relative absence of ground-level refuges. Further research assessing habitat quality and local threats is required to explain the relative absence of threatened herpetofauna from golf courses in SEQ. If the low diversity of threatened reptiles and amphibians does simply reflect a lack of habitat complexity, there are likely to be substantial opportunities to improve the quality of herpetofaunal habitats on golf courses, particularly at a time where many golf clubs are looking to streamline their maintenance costs by reducing the area of land that is maintained (eg. Pearce, 1993).
31 2.6.3 THE ECOLOGICAL ROLE OF HABITAT ON GOLF COURSES In theory, well-vegetated golf courses could act as a supply of additional resources, as a relatively benign land type that could buffer reserves from more threatening urban land types, as a source of connecting habitat or as a localised source of independently productive habitat. Habitats on golf courses will inevitably perform different ecological roles for different animals. Species-specific studies would therefore be required to define the ecological role played by habitats on golf courses. In addition, more information is required on the nature and intensity of local threats faced by wildlife. While observations made during sampling indicate that the threats of predation, disturbance, noise, traffic and death by collision may be lower on golf courses than in other urban areas, the threat of chronic, low-level exposure to chemical herbicides remains a credible threat that warrants research attention. While this study has shown that threatened vertebrates can persist on well- vegetated golf courses, the local densities observed are not necessarily an indication of local reproductive success (eg. Van Horne, 1983; Vickery et al., 1992). Wildlife movements can redistribute animals throughout the landscape, thereby elevating species richness in sub-optimal habitats (Pulliam, 1988; Pulliam and Danielson, 1991; Dunning et al., 1992). High biodiversity levels observed on some golf courses may therefore be to some extent, an artefact of historical or contextual influences. Population ecology studies, assessing local productivity and survivorship could predict the long-term viability of wildlife populations on isolated golf courses and determine whether golf courses should be actively connected to formal reserve networks, given the potential risks and benefits associated with connecting potentially suboptimal (i.e. sinks) and optimal habitat (i.e. sources; Lidicker, 1975; Van Horne, 1983; Pulliam, 1988; Pulliam and Danielson, 1991; Delibes et al., 2001). While many questions remain, this chapter has shown that golf courses are not too small to retain habitat that has value for a range of regionally threatened vertebrates. However, while some golf courses in SEQ clearly have the capacity to make a positive, albeit localised contribution to urban wildlife conservation by restricting the localised decline of species that are regionally disappearing due to urbanisation, many others have no conservation value whatsoever. It is important to determine the source of this variation and in particular, the extent to which it is attributed to golf course design and management practices, as these represent practical mechanisms for ecological enhancement on golf courses. Golf course design and management practices are however, unlikely to be the sole
32 determinant of local biodiversity levels on golf courses. Local biodiversity is rarely a simple reflection of local habitat size and quality (Marzluff and Ewing, 2001). Wildlife assemblages are shaped by dynamic biotic-abiotic associations and ecological interactions operating at multiple spatial and temporal scales (Wiens, 1989; George and Zack, 2001). While efforts to increase the size and complexity of habitats on golf courses may have some impact on local wildlife assemblages, other regional and temporal factors may also affect biodiversity and therefore restrict the potential for ecological enhancement. Factors affecting biodiversity on golf courses in SEQ will be investigated (Chapter 6), to determine the potential for small-scale conservation and restoration within the golf industry. It is also essential to develop a better understanding of the species occurring on individual urban land types. While golf courses provided habitat for some refuge-dependent species, they failed to accommodate others. Information on the ecological characteristics distinguishing species that can be accommodated on small urban habitats (such as golf courses), from those that require larger reserves is required. This will provide greater understanding of the ecological value and inherent limitations of small habitat remnants. This will be investigated in the next chapter.
33
34 CHAPTER 3 ECOLOGICAL CHARACTERISTICS OF WINNERS AND LOSERS
3.1 INTRODUCTION Wildlife species are not equally disadvantaged by urbanisation (Harrison, 1993; Vitousek et al., 1996; Baskin, 1998; McKinney and Lockwood, 1999; Low, 2002). While many species experience localised declines and extinctions, some actually benefit from urban land modification (McKinney and Lockwood, 1999; Blair, 2001; Low, 2002). Urbanisation therefore represents an intense selection pressure that is imposed on wildlife communities around the world. This pressure acts at a localised level, transforming existing wildlife assemblages, creating a predictable suite of ‘winners’ and ‘losers’ (Blair, 2001). In this way, urban sprawl gradually facilitates the range expansion of winners and the contraction of losers (McKinney and Lockwood, 1999). At a regional level this results in a significant loss of species diversity in a process commonly termed ‘biotic homogenisation’ (McKinney and Lockwood, 1999). To develop comprehensive, long-term urban conservation strategies it is important to understand the level of threat faced by individual species; to effectively distinguish winners from losers. It is also essential to identify species that can be effectively protected in urban conservation efforts. Because species vary in their vulnerability to habitat fragmentation, some are more amenable to small-scale conservation efforts than others. An understanding of the ecological requirements of individual species is required to ensure conservation efforts have ecologically realistic goals and target the appropriate species (Ehrenfeld, 2000). This is particularly important for smaller-scale conservation efforts. While the previous chapter has shown small habitats (i.e. those retained on golf courses) can support refuge-dependent, (suburban avoiding) wildlife, they are unlikely to accommodate all threatened species. It would be useful to investigate the ecological characteristics that distinguish winners from losers and species that can be supported on small habitats from those that require larger reserves. This would help define the conservation role that could be played by small semi-natural habitat remnants. Certain ecological characteristics predispose species to relative success or failure in fragmented landscapes (Bentley et al., 2000; Owens and Bennett, 2000; Purvis et al., 2000; Henle et al., 2004). Studies investigating wildlife responses to habitat fragmentation have identified ecological attributes that heighten species vulnerability to
35 landscape change. These include natural rarity (Soulé et al., 1988; Bolger et al., 1997), intermediate-large body size (Kitchener et al., 1982; Burbridge and McKenzie, 1989; Karr, 1990; Gaston and Blackburn, 1995), low adult survival rates (Karr, 1982), low fecundity (MacArthur and Wilson, 1967; Sieving and Karr, 1997), high trophic level (Crooks and Soulé, 1999), low tolerance of the matrix (Brown and Kodric-Brown, 1977; Lomolino, 1986; Laurance, 1991; Sarre et al., 1995), large home range (Woodroffe and Ginsberg, 1998; Webb and Shine, 1998) and ecological specialisation (Patterson, 1987; Foufopoulos and Ives, 1999; Bentley et al., 2000; Ford et al., 2001). The ecological traits associated with local species decline are likely to vary depending on the underlying ecological mechanisms responsible (Owens and Bennett, 2000). This chapter investigates the ecological characteristics that distinguish winners from losers among vertebrate wildlife in urban areas of SEQ and characteristics that distinguish species that can be accommodated on small urban remnants (i.e. golf courses), from those that require larger remnants of native eucalypt habitat. 3.2 AIMS This chapter has two specific aims: 1) to identify life history characteristics that distinguish winners from losers 2) to distinguish groups of urban-threatened wildlife that can be supported on high refuge golf courses from those that are abundant only in forest areas 3.3 METHODS The results of fauna surveys (Chapter 2) were used to assess the prevalence of individuals with shared ecological characteristics in each land type (residential areas, golf courses and eucalypt fragments). For each species found in surveys, ecological characteristics were researched from published literature. Species were then assigned to ecological groupings based on shared ecological and life history characteristics. Birds were separated into ecological groups based on: 1) foraging mode (low granivores, low omnivores, hawk-gleaners, pounce-gleaners, high gleaners) 2) microhabitat (cleared land, open mid-storey, complex understorey, closed mid-storey, canopy) 3) body size (0-10cm, 11-20cm, 21-30cm, 31-40cm, 41-50cm) 4) territoriality and aggression (breeding territories, permanent territories, aggressive) and 5) migrant status (forest resident, forest migrant). Reptiles were categorized by: 1) microhabitat (semi-aquatic, simple ground cover, arboreal, fossorial, complex ground cover) and 2) taxonomic groupings (Agamids, Snakes, Skinks, Varanids). Mammals were categorized according to: 1) home range size (< 1ha, 1-2ha, 2-10ha, >10ha) and 2) functional
36 groups (exotic rodents, native rodents, urban-tolerant arboreal, urban-avoiding arboreal, macropods). Amphibians were separated into ecological groups based on: 1) breeding site (generalist, permanent waterbodies, streams, ephemeral waterbodies) 2) microhabitat preferences (terrestrial, aquatic vegetation, aquatic, edge vegetation, arboreal), 3) larval duration (<50 days. 50-90 days, 100-200 days, >200 days) and 4) fecundity/clutch size (0-250 eggs, 250-1000, 1000-10000, >10000 eggs). References used to research life history characteristics are shown in Appendices 2-5. Key base references included (Wade, 1975; Frith, 1976; Slater et al., 1989; Pizzey and Knight, 1999) – for birds, (Cogger, 1992; Anstis, 2002) – for reptiles and amphibians and (Strahan, 1998) – for mammals. 3.4 DATA ANALYSIS The cumulative site abundance of species in each ecological grouping was calculated from the results of fauna surveys (Chapter 2). The abundance of individuals in each group was compared among residential areas, golf courses and eucalypt forests. Pairwise comparisons were made between each land type using Mann-Whitney U-tests to distinguish; 1) winners from losers (i.e. comparing suburban areas with eucalypt forests) 2) species supported on golf courses from those that require interior habitat (i.e. comparing ‘high-value’ golf courses with eucalypt forests) Bonferroni corrections were used to adjust 0.05 significance levels. Power analyses were used to show the reliability of all non-significant results. 3.5 RESULTS Birds – Microhabitat Selection Birds that were abundant in eucalypt fragments were distinguished from those in residential areas and on most golf courses by ecological characteristics including foraging mode, nest height, territoriality, body size and microhabitat preferences. Birds that nested in closed canopy, closed mid-storey and understorey vegetation were significantly less abundant in residential areas ( x closed canopy = 26, p < 0.0001; x closed mid- storey = 10.7, p = 0.0002; x understorey = 1.1, p < 0.0001) than in eucalypt remnants ( x closed canopy = 103, x closed mid-storey = 45.3, x understorey = 47.9; Fig. 3.1). These birds were however found in relatively high densities on some golf courses. Golf courses supported significantly fewer understorey nesting birds ( x = 18.8) than eucalypt fragments ( x = 47.9, p = 0.002), but significantly more than suburban areas ( x = 1.1, p = 0.003). Three golf courses supported higher local densities of understorey and
37 canopy-nesting birds than were found on average, in eucalypt forests (Fig. 3.1). Golf courses supported numbers of closed mid-storey birds ( x = 27.8) that were neither significantly lower than those found in eucalypt fragments ( x = 45.3; Power = 0.51), nor significantly higher than in residential areas ( x = 10.7; Power = 0.62). Birds that are typically adapted to life in cleared land and open mid-storey vegetation were significantly more abundant in residential areas ( x cleared = 145.8, x open mid = 171.4) than in eucalypt forest ( x cleared = 39.4, p < 0.0001; x open mid = 99.5, p = 0.01). Golf courses supported more urban-adapted (i.e. cleared land and open mid-storey) birds than eucalypt forests, but fewer cleared land birds than residential areas (p < 0.0001). Open mid-storey birds were found in comparable abundance on both golf courses and in residential areas. There was however insufficient power to detect a significant difference in this instance (Power = 0.08). Birds – Body size Smaller birds (<10cm and 11-20cm) were significantly less abundant in residential areas ( x <10cm = 7.8, p < 0.0001, x 11-20cm = 21.5, p = 0.0003) than in eucalypt remnants
( x <10cm = 61.7, x 11-20cm = 75) and are therefore likely to be adversely affected by urbanisation (Fig. 3.1). Golf courses supported small urban-threatened birds (sized 11- 20cm) in numbers ( x = 45.6) comparable to those found in eucalypt forests (Power = 0.63). Very small birds (<10cm) were however significantly less abundant on golf courses ( x = 16.1) than in forests ( x = 61.7, p = 0.0009). Mid-sized birds (21-30cm) were abundant in both residential areas and golf courses. Birds – Diet and Foraging strategy Diet and feeding strategies that were shared by birds that were less abundant in residential areas and on golf courses than in eucalypt remnants (and which are therefore likely to be threatened by urbanisation) included hawk gleaners ( x eucalypt = 54.1, x suburb
= 7.3, p = 0.0006), insectivores ( x eucalypt = 105.8, x suburb = 33.7, p < 0.0001; x golf =
23.1, p = 0.01) and pounce gleaners ( x eucalypt = 76.7, x suburb = 48.5, p < 0.0001; x golf = 55.5, p = 0.003; Fig. 3.1). Among these, golf courses supported significantly more hawk gleaners and pounce gleaners than residential areas. Low granivores were significantly more abundant in residential areas ( x = 72.5) than in eucalypt forest ( x = 19.4, p < 0.0001) and golf courses ( x = 31.8, p = 0.0002). Low omnivores were significantly more abundant in residential areas ( x = 92.2, p < 0.0001) and golf courses ( x = 92.2, p < 0.0001) than in eucalypt forest ( x = 40.9).
38 Birds – Migrant status Both forest migrants and residents were significantly more abundant in eucalypt forests
( x res = 42.8, x mig = 90) than on golf courses (pres = 0.002, pmig < 0.0001) and suburban areas (pres < 0.0001, pmig < 0.0001). Forest residents were significantly more abundant on golf courses ( x = 16.4) than in residential areas ( x = 1.8, p = 0.007). While there was no difference in the abundance of forest migrants between golf courses ( x = 20.5) and suburban areas ( x = 7.6; Fig. 3.1), the power of this test was relatively low (0.47) and may have prevented the detection of a significant difference. Birds – Territoriality and Aggression Birds that maintain permanent territories tended to be significantly more abundant in modified landscapes (i.e. suburban areas x = 146.8, and golf courses x = 144.5) than in eucalypt forest ( x = 78.7, psub = 0.0003, pgolf = 0.0006). Similarly, aggressive birds were significantly more abundant in residential areas ( x = 169.7) than in eucalypt forest ( x = 115.2, p 0.03). In contrast, birds that maintain breeding territories were significantly more abundant in eucalypt forest ( x = 88.9) than in suburbia ( x = 22.2, p < 0.0001) and on golf courses ( x = 39.3, p = 0.0006; Fig. 3.1). Amphibians Amphibians that were abundant in eucalypt fragments were distinguished from those commonly found on golf courses by clutch size, as well as breeding site and microhabitat preferences (Fig. 3.2). These factors are likely to be correlated. Amphibians that were poorly represented on golf courses, compared with eucalypt fragments included those with small clutches (0-250 eggs, x euc = 130.2, x golf = 12.6, p
= 0.0005), amphibians that typically breed in ephemeral water-bodies ( x euc = 159.7,
x golf = 10.8, p = 0.0002) and those typically associated with sedge habitats ( x euc =
127.4, x golf = 2.15, p = 0.0004; Fig. 3.2). Amphibians that were significantly more abundant on golf courses than in eucalypt forests included species with large clutch sizes (>10,000 eggs, essentially one species: Bufo marinus; x golf = 48.2, x euc = 22.9, p
= 0.0008), amphibians that breed in permanent ponds ( x golf = 64.1, x euc = 12.2, p = 0.0002) and those typically associated with aquatic vegetation, essentially one species;
Litoria fallax ( x golf = 60.9, x euc = 10.9, p = 0.0002).
39 SUBURBIA GOLF COURSES EUCALYPT FOREST
300
150 Abundance Abundance
0 Cleared Open Under- Closed Canopy Cleared Open Under- Closed Canopy Cleared Open Under- Closed Canopy Land Midstorey storey Midstorey Land Midstorey storey Midstorey Land Midstorey storey Midstorey Nesting site
SUBURBIA GOLF COURSES 240 EUCALYPT FOREST
180
120
Abundance Abundance 60
0 Low Low Pounce Hawk High Low Low Pounce Hawk High Low Low Pounce Hawk High Granivore Omnivore Gleaner Gleaner Gleaner Granivore Omnivore Gleaner Gleaner Gleaner Granivore Omnivore Gleaner Gleaner Gleaner Feeding strategy
SUBURBIA 400 GOLF COURSES EUCALYPT FOREST
200 Abundance Abundance
0 0-10cm 11-20cm 21-30cm 31-40cm 41-50cm 0-10cm 11-20cm 21-30cm 31-40cm 41-50cm 0-10cm 11-20cm 21-30cm 31-40cm 41-50cm Body size
240 SUBURBIA GOLF COURSES EUCALYPT FOREST
180
120
Abundance Abundance 60
0 Permanent Breeding Forest Forest Permanent Breeding Forest Forest Permanent Breeding Forest Forest Territories Territories Residents Migrants Territories Territories Residents Migrants Territories Territories Residents Migrants Territoriality and Migrant Status
Figure 3.1. Total site abundance of birds with different ecological characteristics in residential areas, golf courses and eucalypt forests.
40 GOLF COURSES EUCALYPT FORESTS
240 240
180 180
120 120
Abundance 60 60
0 0 0-250 250-1000 1000-10000 >10000 0-250 250-1000 1000-10000 >10000 Clutch size Clutch size
240 240
180 180
120 120
Abundance 60 60
0 0 <50 50-90 100-200 >200 <50 50-90 100-200 >200 Larval duration Larval duration
300 300
200 200
100 100 Abundance
0 0 Generalist Permanent Streams Ephemeral Generalist Permanent Streams Ephemeral Ponds Ponds Ponds Ponds Breeding site Breeding site
400 400
200 200 Abundance
0 0 Terrestrial Aquatic Aquatic Edge Arboreal Terrestrial Aquatic Aquatic Edge Arboreal Veg Veg Veg Veg Veg Veg
Microhabitat Microhabitat
Figure 3.2. Total site abundance of amphibians with different ecological characteristics on golf courses and eucalypt forests.
41 GOLF COURSES EUCALYPT FORESTS
1.8 1.8
1.2 1.2
0.6 0.6 Abundance
0.0 0.0 Exotic Native Common Rare Macropod Exotic Native Common Rare Macropod Rodent Rodent Arboreal Arboreal Rodent Rodent Arboreal Arboreal Functional groups Functional groups
1.8 1.8
1.2 1.2
Abundance 0.6 0.6
0.0 0.0 <1ha 1-2ha 2-10ha >10ha <1ha 1-2ha 2-10ha >10ha Home range Home range Figure 3.3. Total site abundance of different mammal groups
120 120
80 60
40 Abundance
0 0 Aquatic Open Arboreal Fossorial Complex Aquatic Open Arboreal Fossorial Complex Understorey Understorey Understorey Understorey A Microhabitat Microhabitat
240 240
180 180
120 120 Abundance 60 60
0 0 Agamids Snakes Skinks Varanids Agamids Snakes Skinks Varanids Taxonomic Groups Taxonomic Groups
Figure 3.4. Total site abundance of different reptile groups
42 Mammals Few ecological characteristics distinguished differences in the abundance of mammals between golf courses and eucalypt forests. Home range had no significant association with relative abundance in either land type. Small suburban tolerant ground mammals were significantly more abundant on golf courses ( x = 0.5) than in eucalypt forests ( x = 0.05, p < 0.0001; Fig. 3.3). Mammals with high fecundity (>10 young per year) were also significantly more abundant on golf courses ( x = 2.3) than in eucalypt forest ( x = 0.2, p = 0.0001). Suburban-intolerant arboreal mammals were significantly more abundant in eucalypt forest ( x = 0.3) than on golf courses ( x = 0.1, p = 0.01). Reptiles Reptiles significantly more abundant on golf courses than in eucalypt forests included semi-aquatic species ( x golf = 21.8, x euc = 1.1, p = 0.003) and Agamids ( x golf = 22.8,
x euc = 2.7, p = 0.003). Both of these effects were attributed to one species, the eastern water dragon, Physignathus lesueurii which is commonly found close to permanent waterbodies on golf courses. Reptiles that were poorly represented on golf courses compared to eucalypt forest included species typically associated with complex ground cover ( x euc = 15.9, x golf = 2.2, p = 0.0003) and Varanids (essentially Varanus varius,
x euc = 2.9, x golf = 0.4, p = 0.02; Fig.3.4). Snakes were rarely encountered in both eucalypt forests and golf courses. Greater sampling effort would be required to compare their relative abundance between the two land types. 3.6 DISCUSSION Wildlife that tolerate or exploit urban landscapes typically include ecological generalists and species pre-adapted to urban conditions (i.e. cleared land and edge specialists). In contrast, species with specialist diet or microhabitat requirements tend to be restricted to natural habitats (Terborgh and Winter, 1980; Patterson, 1987; Laurance, 1990; 1991; Collinge, 1996; Bentley and Catterall, 1997). The results of this study supported that trend. In all cases, habitat and diet generalists were as abundant on golf courses and residential areas as in remnants of eucalypt forest. In contrast, habitat and diet specialists were largely restricted to eucalypt forests.
3.6.1 BIRDS Diet, body size, feeding behaviour and microhabitat selection were useful traits distinguishing the relative abundance of birds in each land type. Similar results have been observed in other studies, suggesting bird habitat selection is a function of habitat structure and foraging/nesting behaviour (Johnston and Odum, 1956; Blake, 1982; Ding
43 et al., 1997). Low granivores, omnivores and species typically associated with cleared land were significantly more abundant in residential areas (and to a lesser extent golf courses) than in eucalypt forests. Low granivores and omnivores benefit from urbanisation in many parts of the world (Emlen, 1974; Lancaster and Rees, 1979; Beissinger and Osborne, 1982; Mills et al., 1989; Catterall et al., 1998; Lim and Sodhi, 2004). Their success is attributed to the increased availability of grass seeds, ground insects and discarded food (Emlen, 1974). In Australia, urban areas have been invaded by arid-zone granivores including the crested pigeon and galah (Catterall and Kingston, 1993). Similar range expansions have been observed among arid-land granivores in the United States (Emlen, 1974; Mills et al., 1989). The relatively low abundance of low granivores on golf courses could be attributed to the use of low-seeding varieties of turf-grass and a relative lack of suitable nesting sites. Many urban-adapted granivores nest and roost in the eaves of buildings and other man-made structures that are not present on golf courses. Large, aggressive nectarivores and frugivores were significantly more abundant in residential areas and golf courses than in eucalypt forest. Exploitation of suburban habitats by large nectarivores has been documented elsewhere in Australia (Green, 1984; Wood, 1996; Sewell and Catterall, 1998; Parsons et al., 2003; French et al., 2005) and in subtropical and tropical urban environments around the world (Petit et al., 1999; Lim and Sodhi, 2004). The success of large nectarivores is attributed to urban ornamental plantings that are often high in nectar content (Grey et al., 1997; Pell and Tideman, 1997). This increased food supplement has enabled birds that were traditionally nomadic feeders (McGoldrick and MacNally, 1998) to maintain permanent suburban territories, which they defend aggressively (French et al., 2005). Small insectivores were significantly more abundant in eucalypt forests than on golf courses and residential areas. Small insectivores and honeyeaters are universal losers in urban landscapes throughout the world and are often particularly vulnerable to edge effects (Emlen, 1974; Beissinger and Osborne, 1982; Catterall et al., 1997; 1998; Clergeau et al., 1998; Rottenborn, 1999; Allen and O’Conner, 2000; Lindsay et al., 2002; Lim and Sodhi, 2004). While edge-sensitivity is common among small insectivores, the mechanisms driving the response vary regionally. In North America, it may be caused by increased predation and brood parasitism (Brittingham and Temple, 1983; Bohning-Gaese et al., 1993; Robinson and Wilcove, 1994). However, in Australia it is predominantly attributed to competitive exclusion by aggressive
44 nectarivores, particularly the noisy miner (Manorina melanocephala; Catterall et al., 1997; Grey et al., 1997; MacNally et al., 2000; Major et al., 2001; French et al., 2005). The relative absence of small-insectivores from residential areas and many golf courses may also be associated with the loss and simplification of local habitats in these land types. Small insectivores typically have specialist feeding and nesting habitat requirements and are therefore unlikely to persist in areas where habitat structure has been substantially degraded (Barrett et al., 1994; Ford et al., 2001).
3.6.2 MAMMALS Large mammals can be particularly vulnerable to urban and agricultural habitat fragmentation (Bruinderink et al., 2003). In SEQ, large macropods such as the eastern grey kangaroo (Macropus giganteus), red-necked wallaby (Macropus rufogriseus) and swamp wallaby (Wallabia bicolor) do not generally persist in suburban areas. However, while these species have relatively large area requirements (Grigg et al., 1989) and are therefore unlikely to persist in small remnants, there was no significant difference in macropod abundance between golf courses and eucalypt remnants. The failure to detect a significant difference in macropod abundance between golf courses and eucalypt fragments does however, not necessarily sugest that golf courses are a comparable source of macropod habitat to eucalypt bushland. High variation in macropod abundance among golf courses meant the analysis had insufficient power (0.26) to detect a significant difference. Four of the eucalypt remnants used for comparison were relatively small (<200ha) and may therefore not adequately represent ideal macropod habitat. In addition, macropods may have been more easily detected on golf courses than in eucalypt forests (where hiding refuges are more abundant). Nevertheless, macropods evidently persist in substantial numbers on a number of golf courses in SEQ. The long-term viability of these populations may depend on the level of connectivity that is maintained to external habitats. Small macropod populations may be unlikely to persist indefinitely on isolated golf courses. In urban landscapes, the fate of smaller mammals depends on ecological characteristics including habitat specialisation, matrix tolerance, dietary specialisation and reproductive capacity (Dickman and Doncaster, 1989). Small mammals that persist in urbanised habitats (eg. the black rat and house mouse) generally have broad ecological tolerances, high fecundity and matrix tolerance (Dickman and Doncaster, 1989). In this study, small ground mammals with high matrix tolerance and high fecundity (i.e. black rats and house mice) were significantly more abundant on golf
45 courses than in eucalypt forests. The relative absence of the house mouse (Mus musculus) from native remnants is typically attributed to competitive exclusion by larger native rodents (Haering and Fox, 1993; Bentley et al., 2000). The relative absence of the black rat (Rattus rattus) from native remnants has been attributed to both competitive exclusion (Diffendorfer et al., 1995) and habitat preferences (Watts, 1995; Dunstan and Fox 1996; Bentley et al., 2000). In this study, refuge-dependent native rodents were found in comparable abundance on both golf courses and in eucalypt forests. The failure to detect a significant difference may however, be attributed to the relatively low power (0.47) of the statistical test. Greater trapping effort would be required to reliably determine if golf courses are indeed supporting comparable densities of regionally threatened rodents to those found in larger eucalypt forests.
3.6.3 REPTILES Reptiles and amphibians have fundamentally different ecological requirements to other vertebrates. Herpetofauna have strict thermal and body-moisture constraints (Duellman and Trueb, 1986; Heatwole and Taylor, 1987). Their behaviour, movement and use of habitats are therefore strongly influenced by environmental and physiological conditions (Duellman and Trueb, 1986; Heatwole and Taylor, 1987). Many reptiles and amphibians have limited mobility compared to other vertebrates (Fitch and Shirer, 1971; Harris, 1975; Beshkov and Jameson, 1980; Blaustein et al., 1994) and respond to the landscape at a relatively fine scale, often having close associations with specific microhabitats. Ground-level vegetation is often structurally simplified in urban areas. This has dramatic consequences for local reptile assemblages. Reptiles that persist in suburban landscapes tend to include species with broad microhabitat tolerances or species specifically adapted to simple ground cover (Germaine and Wakeling, 2001). In this study, reptiles typically associated with complex ground cover were significantly more abundant in eucalypt forest than on golf courses. In contrast, species associated with simple ground cover were abundant in both land types. Semi-aquatic reptiles including the eastern water dragon (Physignathus lesueurii) and eastern water skink (Eulamprus quoyii) were significantly more abundant on golf courses than in eucalypt forests. These species are evidently well-adapted to the conditions provided by artificial permanent ponds. Another reptile that tolerated the environments on golf courses and which is expected to be even more abundant in residential areas is the wall skink (Cryptoblepharus virgatus). Adapted to microhabitats on tree trunks, this species is abundant in wood piles on golf courses and on fences,
46 walls and other flat vertical substrates in suburbia. A species with a similar ecological niche, the tree lizard (Urosaurus ornatus) has benefited from urbanisation in Arizona, USA (Germaine and Wakeling, 2001). The proliferation of vertical substrates in urban landscapes may provide a useful ecological niche for reptiles that can forage vertically. In south-east Queensland this niche is also used nocturnally by the Asian House Gecko (Hemydactylus frenatus).
3.6.4 AMPHIBIANS Amphibian breeding site preferences (and associated ecological traits) were useful in distinguishing amphibians capable of persisting on golf courses from those that require larger remnants of natural eucalypt habitat. The apparent association between land-use and breeding site selection was not unusual. Many amphibian species are specifically adapted to breed in certain water-body types, having evolved distinct strategies to overcome the inherent threats (of predation and environmental stress) that are unique to particular water-bodies (Wiggins et al., 1980; Schneider and Frost, 1996; Wellborn et al., 1996; Hero et al., 2001). Species that breed in permanent ponds are exposed to high predation pressures and have therefore evolved anti-predator defences (i.e. crypsis and unpalatability; Semlitsch and Gavasso, 1992; Semlitsch and Reyer, 1992; Snodgrass et al., 2000). In contrast, species that breed in semi-permanent and ephemeral water- bodies experience greater threat from periodic desiccation (Lannoo, 1998; Semlitsch and Bodie, 1998; Snodgrass et al., 2000) and therefore have few anti-predator defences, but forage actively and metamorphose quickly (Wellborn et al., 1996). Amphibians have little capacity to switch strategies (Wiggins et al., 1980; Schneider and Frost, 1996; Wellborn et al., 1996) and are therefore vulnerable to changes in local hydrology (Beebee, 1997; Hazell et al., 2003; Rubbo and Kiesecker, 2005). Urban and agricultural landscape changes tend to favour increased waterbody permanence, as ephemeral and semi-permanent waterbodies are replaced by permanent ponds and dams (Hazell et al., 2003; Rubbo and Kiesecker, 2005). This selectively disadvantages species dependent on semi-permanent and ephemeral breeding sites. This situation also occurs on golf courses, where there is a tendency to build or retain permanent ponds (for aesthetic, irrigation and flood-water mitigation reasons) and reduce the number of temporary waterbodies (to provide a constant, predictable environment for golfers). The capacity for amphibians to tolerate urban/agricultural landscape changes is also partly influenced by their dispersal ability and matrix tolerance. These
47 characteristics determine the extent to which populations will become isolated by habitat fragmentation. Contrary to traditional perceptions, amphibians can be susceptible to the effects of habitat fragmentation (Laan and Verboom, 1990; Sjögren, 1991; Blaustein et al., 1994; Marsh and Pearman, 1997). Many species (particularly those that breed in isolated permanent ponds) are dependent on ongoing dispersal and recolonisation in order to persist locally (Sinsch, 1997; Seburn et al., 1997; Semlitsch, 2000) and are therefore particularly vulnerable to isolation effects (Green, 2003). Species adapted to modified landscapes in Europe (eg. Bufo bufo, Rana esculenta, Rana temporaria and Hyla intermedia) tend to have high dispersal abilities, high matrix tolerance and favourable demographics (Green, 2003; Ficetola and De Bernardi, 2004). In this study, one species with high fecundity and matrix tolerance (Bufo marinus) was significantly more abundant on golf courses than in eucalypt forests. The cane toad has been particularly successful in modified landscapes and owes its success to a combination of broad ecological tolerance, rapid development, high dispersal ability and very high fecundity (Easteal and Floyd, 1986; Lever, 2003).
3.6.5 CHARACTERISTICS OF WILDLIFE ON GOLF COURSES In general, most golf courses were utilised by species with broad ecological tolerances and species adapted to modified or edge-affected habitats. Species that are dependent on complex mid-level and understorey vegetation, complex ground cover and ephemeral waterbodies were significantly less abundant on most golf courses than in eucalypt forests. Only a small number of golf courses retained ecological specialists that are threatened by urbanisation. Understanding the conditions that have allowed threatened species to persist on some golf courses and not others is essential if the conservation potential of suburban golf courses and urban ‘open-space’ areas is to be realised. It is important to determine if golf course design and management practices can influence the conservation value of individual golf courses, or if the high diversity of threatened vertebrates recorded on some golf courses is attributed to other historic or regional environmental influences. These questions will be answered in Chapter 6.
48 CHAPTER 4 A REVIEW OF LANDSCAPE ECOLOGY THEORY
INTRODUCTION Concern for the welfare of wildlife has inspired research investigating the relationship between landscape structure and local biodiversity. Landscape ecology is a new multi- disciplinary science that has evolved to address the ecological problems created by human land-use and develop spatially explicit conservation strategies (Forman and Godron, 1986). This chapter presents a review of landscape ecology theory to highlight the extent of current knowledge. While landscape ecology is a relatively new discipline, it is based on ecological principles that have been established for some time. The theories of island biogeography (MacArthur and Wilson, 1963; 1967) and metapopulation dynamics (Levins, 1969) provide a conceptual framework by which to understand the response of wildlife communities in fragmented landscapes. THE THEORY OF ISLAND BIOGEOGRAPHY The theory of island biogeography was initially proposed to explain the composition of wildlife communities on oceanic islands (MacArthur and Wilson, 1963) but was later applied to ‘islands’ of terrestrial habitat, isolated by areas of inhospitable land (MacArthur and Wilson, 1967). The theory suggests that species richness on an island (or habitat fragment) should be directly related to its size and proximity to a source of colonising species, as these factors ultimately determine local rates of immigration and extinction (MacArthur and Wilson, 1963). Colonisation rates are higher on islands that are relatively close, while rates of extinction are lower on islands that are large (MacArthur and Wilson, 1963). As a result, species richness should be higher on large, close islands than those that are small and remote (MacArthur and Wilson, 1963; 1967). THE THEORY OF METAPOPULATION DYNAMICS The theory of metapopulation dynamics was initially proposed to predict the population dynamics of species occurring in naturally fragmented habitats (Levins, 1969). More recently it has been applied to wildlife populations whose habitats have been fragmented by human land-use activities (Murphy et al., 1990; Harrison, 1994; Doak and Mills, 1994; Hanski et al., 1995). According to the theory, a metapopulation consists of a series of spatially independent satellite populations that can experience periodic colonisation and extinction, while the metapopulation as a whole continues to
49 persist (Bleich et al., 1990; Hanski and Gilpin, 1991). Ongoing movement of individuals between sub-populations is critical for metapopulation persistence (Harrison and Quin, 1989; Hanski, 1989; Olivieri et al., 1990). HISTORY OF LANDSCAPE ECOLOGY RESEARCH While these theories are central to our interpretation of landscapes and the underlying ecological mechanisms governing the distribution and abundance of wildlife, there has been intense debate over the extent to which they can be practically applied to conservation reserve design. Early studies that applied the theory of island biogeography to the design of nature reserves suggested that a single large reserve should be superior to a network of smaller reserves (Diamond, 1975; Terborgh, 1975; Wilson and Willis, 1975). This interpretation was however, strongly criticised by some ecologists who suggested that there would be circumstances where a series of well- placed, smaller reserves may have greater conservation value than a single large reserve (Simberloff, 1976; Simberloff and Abele, 1976; 1982; Gilbert, 1980). This became known in the literature as the ‘SLOSS’ debate (i.e. Single Large Or Several Small). The SLOSS debate was directed not only by ecological theory, but by social considerations that influenced the practical implementation of conservation strategies. Proponents of large reserves (Diamond, 1976; Terborgh, 1976; Whitcomb et al., 1976) warned that the argument to protect smaller reserves may present developers with a biologically defensible rationale by which to justify the ongoing fragmentation of natural habitats (Lynch, 1987; Loyn, 1987). In contrast, proponents of smaller reserves suggested that many conservation opportunities were being overlooked by only recognising value in larger reserves (Primack, 1993). While the SLOSS debate remains largely unresolved, ecologists concede that there is merit in both arguments. Some suggest the debate is irrelevant in urban areas where social and economic constraints rarely present an opportunity to design ‘optimal’ conservation reserves (Simberloff, 1988; Franklin, 1993). A case-by-case approach to conservation is therefore often recommended (Game and Peterken, 1984; Soulé and Simberloff, 1986). PATCH SIZE The SLOSS debate has inadvertently drawn attention to the influence that patch size can have on local biodiversity. In general, large habitat patches support higher species richness than patches that are relatively small (Burgman and Lindenmayer, 1998). Similar conclusions were independently derived by Preston (1962), and other researchers who investigated the species-area relationship, a phenomenon whereby
50 species richness increases with patch size (Moore and Hooper, 1975; Rafe et al., 1985; Tilghman, 1987; Loyn, 1987; Kerley et al., 1996; Major et al., 2001; Vanhinsbergh et al., 2001; Drinnan, 2005). The relationship between patch size and species richness has lead some ecologists to suggest the existence of distinct minimum patch size thresholds, below which habitat will have limited value for native wildlife (Howe et al., 1981; Loney and Hobbs, 1991; Edenius and Sjöberg, 1997; Sewell and Catterall, 1998; Zanette, 2000). Others believe the response to patch size varies and that it is therefore inadvisable to derive a generalised size rule by which to predict local conservation value (Saunders and Hobbs, 1991; Soulé and Gilpin, 1991; Lemckert, 1999). Nevertheless, it is clear that patch size has a significant influence on local species richness. Reasons for the decline in species richness in smaller patches are generally attributed to a decrease in the diversity of available resources and ecological niches and an increase in the relative influence of detrimental ‘edge-effects’. Habitats close to fragment edges are often exposed to higher levels of noise, disturbance, predation, pollution, competition and altered microclimatic conditions (Harris, 1984; Lovejoy et al., 1986; Collinge, 1996). Edge-habitats therefore tend to be relatively degraded and often support fewer threatened species than areas of interior habitat (Whitcomb et al., 1981; Lovejoy et al., 1986; Blake and Karr, 1987; Opdam, 1991). PATCH SHAPE Patch shape has a direct impact on the ratio of edge-to-interior habitat and will therefore affect species diversity. Narrow patches have a greater proportion of edge habitat than rounded patches (Diamond, 1975; Forman and Godron, 1986; Schonewald-Cox and Bayless, 1986) and will therefore tend to support fewer threatened species (Forman and Godron, 1986). The effect of patch shape is however dependent on patch size, having reduced significance as patch size increases (Forman and Godron, 1986). HABITAT COMPLEXITY Local wildlife diversity can be influenced by the structural and floristic complexity of local habitats (Picton, 1979; Simberloff and Gotelli, 1984; Rafe et al., 1985; Boecklen, 1986; Freemark and Merriam, 1986). Evidence to support this idea was first presented by MacArthur and MacArthur, (1961) when they demonstrated that bird species richness could be predicted by foliage height diversity. Subsequent studies have illustrated positive relationships between habitat complexity and local species richness (MacArthur et al., 1962; 1966; Linehan et al., 1967; Recher, 1969; Karr and Roth, 1971; Willson, 1974; Lancaster and Rees, 1979; Beissinger and Osborne, 1982; Rafe et
51 al., 1985; Loyn, 1987; Baines et al., 1994; Young and Armstrong, 1995; Kerley et al., 1996; Gunnarsson, 1996). Different animals respond to different structural elements. While birds often respond to foliage height diversity (Lancaster and Rees, 1979; Beissinger and Osborne, 1982; Humphrey et al., 1999; Bentley et al., 2000), reptiles are typically affected by ground-level indices including shrub-cover, leaf litter thickness, rocky ground and fallen wood (Hadden and Westbrooke, 1996; Smith et al., 1996; Brown, 2001; Jellinek et al., 2004). The abundance of hollow-dwelling mammals and birds can be limited by the abundance of hollow-bearing trees (Meredith, 1984; Smith and Lindenmayer, 1988; Lindenmayer et al., 1990; Traill and Lill, 1990; Smith and Murray, 2003). Many amphibian species respond to aquatic vegetation complexity (Vos and Chardon, 1998; Hazell et al., 2001). While most wildlife respond to variations in habitat structure (Willson, 1974; Terborgh, 1975; Baines et al., 1994; Young and Armstrong, 1995), some (particularly animals with specialist diets) can be affected by variations in floristic diversity (Kitchener et al., 1982; Rowston et al., 2002). Patch size is therefore not the sole determinant of local species diversity. Efforts to enhance the complexity of local habitats may increase local species richness and thereby compensate for the effect of patch size, at least to some extent (Bond, 1957; Lynch and Whigham, 1984). LANDSCAPE CONNECTIVITY Where the theory of island biogeography inspired debate on the ecological significance of patch size, the theory of metapopulation dynamics highlighted the importance of landscape connectivity. Studies have shown that habitat connectivity is a common casualty in urban and agricultural landscapes. This is likely to have a profound impact on wildlife populations, by disrupting their capacity to disperse, migrate or forage throughout the landscape. This has inspired ecologists to consider the possible value in retaining habitat corridors as a way of maintaining landscape connectivity. However, while some ecologists have argued the theoretical value of corridors (Harris, 1984; Forman and Godron, 1986; Noss, 1987; Saunders and de Rebeira, 1991; Saunders and Hobbs, 1991; Dunning et al., 1992; Laurance and Laurance, 1999), others warn that they can act as a conduit for feral predators, competitors, parasites, weeds, disease and fire (Simberloff and Cox, 1987; Noss, 1987; Simberloff et al., 1992). Logistic difficulties have limited research to empirically assess the merits of corridors (Nicholls and Margules, 1991; Lindenmayer and Nix, 1993; Beir and Noss, 1998). Nevertheless, ecologists agree that landscape connectivity is critical to population
52 persistence (Gilpin and Soulé, 1986; Noss, 1987; Primack, 1993; Hunter, 1996). All efforts should therefore be made to protect landscape integrity (Noss, 1987). LANDSCAPE CONTEXT Local biodiversity is rarely a simple reflection of local habitat size and complexity, but is generally determined by multiple factors acting at multiple scales (Turner, 1989; Wiens, 1989; Turner and Gardner, 1991; Dunning et al., 1992; Debinski et al., 2001). Local biodiversity levels can be influenced by the nature and composition of land types within the surrounding regional landscape. Areas of high quality habitat may therefore support fewer species than expected if they are adjacent to land types that are inhospitable or act as a source of harmful influences (Friesen et al., 1995; Rottenborn, 1999). Conversely, areas of relatively low-quality habitat may support high species richness if surrounded by valuable habitat and therefore periodically ‘rescued’ by the re-colonisation of individuals from adjacent productive land (Brown and Kodric- Brown, 1977; Palmer, 1992; Cody, 1993; Holt, 1997). This concept is described in studies of source-sink dynamics and highlights the dangers of assessing habitat quality, simply from patterns of wildlife distribution (Lidicker, 1975; Van Horne, 1983; Pulliam, 1988; Pulliam and Danielson, 1991). Patch occupancy is not necessarily any indication of local habitat productivity (Van Horne, 1983; Vickery et al., 1992). Wildlife movements (both directed and random) can redistribute wildlife throughout the landscape. Areas of productive land (i.e. sources) may elevate species richness in nearby sub-optimal habitats (i.e. sinks). Few studies have investigated the mechanisms driving source-sink dynamics (Dunning et al., 1992). As such, there is limited understanding of the possible risks and benefits associated with connecting areas of source and sink habitat. IMPORTANCE OF THE MATRIX Fragmentation studies have shown that the matrix can have a critical impact on wildlife populations in habitat remnants (Laurance, 1991; Bierregard and Stouffer, 1997; Gascon et al., 1999; Debinski and Holt, 2000). Depending on their nature and juxtaposition, land types in the matrix can act as a selective filter, permitting the movement of some species but restricting others (Laurance, 1997; Gascon et al., 1999). The matrix can also influence the extent to which potentially harmful external factors (i.e. predators, competitors, parasites, disease, noise, disturbance, pesticides) penetrate habitat patches (Laurance and Yensen, 1991; Mesquita et al., 1999; Marzluff and Ewing, 2001). The matrix therefore has the potential to influence populations in
53 isolated habitat patches by acting as a buffer, reducing edge effects and by affecting levels of habitat connectivity (Laurance, 1990; Franklin, 1993; Tocher et al., 1997; Marzluff and Ewing, 2001). Information is therefore required on the nature and influence of all land types in fragmented landscapes. THE EFFECTS OF FRAGMENTATION Fragmentation studies have often produced conflicting results in which the underlying theories of island biogeography and metapopulation dynamics seemingly fail to apply (Debinski and Holt, 2000). Such discrepancies may occur because different processes act and are disrupted at different levels of fragmentation (Franklin and Forman, 1987; Hansen et al., 1992; Collinge, 1996; McIntyre and Hobbs, 1999; Lindenmayer et al., 2003). The ecological effects of fragmentation can be summarised into three broad categories; 1) loss of habitat, 2) reduced fragment size and 3) isolation of remnants (Andrén, 1994). When landscapes are relatively intact, species declines are largely the result of habitat loss (Andrén, 1994). However, at higher levels of fragmentation wildlife are affected primarily by reduced fragment size and by the isolation of remnant habitats (Andrén, 1994; Marsh and Trenham, 2001). When populations become totally isolated, their fate is entirely dependent on local population demographics (Primack, 1993). Small populations will become increasingly susceptible to declines resulting from demographic stochasticity, environmental variation, catastrophic events and genetic drift (Gilpin and Soulé, 1986; Soulé, 1987; Clark and Seebeck, 1990). THE IMPORTANCE OF SPATIAL SCALE The task of assessing the relationship between wildlife communities and the natural landscape is infinitely complicated by the fact that animals vary dramatically in the scale at which they perceive and therefore respond to the land (Wiens, 1976; MacNally et al., 2000). The resolution of wildlife response is determined largely by their morphology, home range and the extent of their seasonal movements (Wiens, 1989; Kotliar and Wiens, 1990). For the purposes of ecological research, landscapes should be assessed from an organism-centred view (Forman and Godron, 1986). As such, there is no single ‘correct’ scale at which to describe the landscape (Greig-Smith, 1964; Steele, 1989; Allen and Starr, 1982; Levin, 1992). The internal structure of any patch therefore reflects patchiness at a finer scale, and a landscape mosaic is characterised by patchiness at broader scales (Kotliar and Wiens, 1990). An understanding of the resolution at which animals respond to the landscape is essential to detect their response and thus to develop spatially explicit conservation strategies (Savard et al., 2000).
54 THE IMPORTANCE OF TEMPORAL SCALE The distribution and abundance of wildlife is however unlikely to be perfectly explained by the current quality and spatial arrangement of habitats in the landscape. Biodiversity observed in the landscape today is partly a function of previous biodiversity levels and the history of local landscape change (Debinski et al., 2001). Landscapes are temporally dynamic. Periodic disturbance and succession are part of normal ecosystem function. Human landscape modifications can alter the frequency of normal ecological processes and introduce new disturbance regimes (Turner, 1989). The loss and fragmentation of habitat caused by agriculture and urbanisation represent a relatively recent phenomenon. Many wildlife communities may still be in the process of responding to those landscape changes. Populations and communities trapped in isolated pockets of habitat are therefore likely to be in some state of decline. Populations are likely to decline at different rates, depending on the animals’ longevity and rate of population turnover (Debinski and Holt, 2000). Animals that have short generation times (eg. insects) are likely to respond more rapidly to landscape change than those that are long-lived (eg. large mammals). CONCLUSION Landscape ecology studies have shown that the relationship between landscape characteristics and local species diversity can be inherently complex and often difficult to classify (Forman and Godron, 1986; Collinge, 1996). Species vary dramatically in their habitat requirements and ecological tolerances and frequently respond differently to landscape characteristics. Designing habitat reserve systems to ‘enhance biodiversity’ is therefore often a difficult process. No single management strategy will benefit all species and almost any habitat enhancement strategy will have negative impacts on some species (Lynch and Whigham, 1984). Nevertheless, biodiversity strategies can often only be approached from a community level. At that scale, simplification is an unfortunate, yet necessary requirement. Landscape ecology studies have been of great use in simplifying the relationship between landscapes and local wildlife communities, and have developed a practical framework for effectively protecting biodiversity. More research is however required to link form with function, to explain the underlying ecological mechanisms responsible for the observed spatial distribution of biota. Research is also required at smaller scales, relevant to the management of habitats in highly fragmented urban and agricultural landscapes.
55
56 CHAPTER 5 ASSESSMENT OF VEGETATION ON GOLF COURSES
5.1 INTRODUCTION Biodiversity on golf courses may be partly determined by the size and quality of habitats retained in rough and out-of-play areas. Landscape ecology studies have found that biodiversity can increase with patch size (Moore and Hooper, 1975; Rafe et al., 1985; Tilghman, 1987; Loyn, 1987), patch roundness (Diamond, 1975; Forman and Godron, 1986), habitat connectivity (Gilpin and Soulé, 1986; Noss, 1987; Hunter, 1996) and with increased structural complexity indices such as foliage height diversity (MacArthur and MacArthur, 1961; Beissinger and Osborne, 1982; Humphrey et al., 1999), ground complexity (Hadden and Westbrooke, 1996; Smith et al., 1996; Brown, 2001; Jellinek et al., 2004), the abundance of fallen wood (MacNally et al., 2001), hollows (Meredith, 1984; Smith and Lindenmayer, 1988; Lindenmayer et al., 1990; Smith and Murray, 2003), aquatic vegetation diversity (Vos and Chardon, 1990; Hazell et al., 2001) and waterbody diversity (Beebee, 1997; Snodgrass et al., 1999; Hazell et al., 2003; Rubbo and Kiesecker, 2004). These factors are likely to affect the diversity of available foraging, nesting, sheltering and breeding opportunities and the nature and intensity of suppressive factors (eg. predation, competition, parasitism and disturbance). The size and structural complexity of local habitats varies among golf courses, depending on their style. Traditional Scottish courses devote approximately 60-65% of their land area to native vegetation (Green and Marshall, 1987; Dair and Schofield, 1990; Brennan, 1992) and typically retain structurally complex vegetation in rough and out-of-play areas. Many traditional Scottish courses have recognised ecological value, containing designated Sites of Special Scientific Interest (SSSI) that are partly managed by the Nature Conservancy Council (NCC, 1989). In contrast, American style golf courses typically retain 10-30% native vegetation, have low structural habitat complexity and have been widely criticised for their detrimental environmental impacts (Pearce, 1993; Smart et al., 1993). There is currently no legislation governing the retention or restoration of vegetation on golf courses in Australia. While American-style course designs have been widely adopted within the Australian golf industry, there is little information on the extent to which Australian golf courses retain native vegetation. This chapter assesses the size and structural complexity of native vegetation retained on twenty representative
57 golf courses in south-east Queensland, (one of Australia’s principal golfing centres). This information will later be used to investigate the potential impact that golf course design and management have on local wildlife assemblages. 5.2 METHODS Spatial characteristics of the terrestrial and aquatic environment were measured at three scales: 1) local - within patches 2) landscape - within golf courses 3) regional - adjacent to golf courses. At each site, local scale (habitat complexity) data were measured on-location, in ten terrestrial and ten aquatic sub-sites, randomly selected from numbered grids overlayed on aerial site-photos. Landscape and regional data were measured using MapInfo GIS, from rectified aerial photographs of each site (Mapview, 2001). 5.2.1 LOCAL-SCALE ENVIRONMENTAL FACTORS In this study, terrestrial ‘patches’ refer to discrete areas of over-storey vegetation, separated from other areas of vegetation by a gap of more than twenty metres (measured directly beneath canopy extremities). The structural complexity of local habitats was measured in all terrestrial patches. Canopy Gap Fraction (CGF) was measured from canopy photos taken at two representative locations in each terrestrial and aquatic patch using methodology outlined in Zancola et al., (2000). Foliage Height Diversity (FHD) was measured at two representative locations in each terrestrial patch using a 2.5m x 0.3m vertical profile board, divided into 50cm height intervals. At each interval, the proportion of the board covered by vegetation was assessed, observing the board from a distance of ten metres. Foliage Height Diversity was calculated using the Shannon-Weaver diversity equation and the mean FHD derived from the two measurements at each sub-site. Ground cover characteristics were measured from photos of a rectangular 0.9m x 0.8m ground cover quadrat that was placed at two representative locations in each patch. Ground cover photos were rectified using a customised coordinate system in MapInfo GIS and used to measure local ground-cover characteristics including the proportion of turfgrass cover, native grass cover, bare ground, woody debris and ground vegetation. In each terrestrial patch, vegetation density measurements were recorded in a 30m x 10m quadrat, counting the number of trees, dead trees, logs and hollows and measuring the circumference of all trees (>5cm) at chest-height (1.4m).
58 Floristic composition was assessed for each terrestrial sub-site and site by classifying all trees in the 30m x 10m quadrat into generic vegetation types: Banksia, Casuarina, Wattle, Melaleuca, Eucalypt (Bloodwood, Ironbark, Smooth-barked Eucalypt), Pine, Palm, Mangrove and Rainforest. In each aquatic sub-site, the relative steepness of waterbody banks (i.e. mean bank angle) and bank height was measured at four points. Ground cover characteristics and CGF were also measured at two locations on the bank of each waterbody, using methods described for terrestrial sub-sites. The proportion of the bank fringed by reed- cover was estimated. For large waterbodies, reed cover was estimated along a 50m length of bank. The proportion of each waterbody occupied by floating, submerged and emergent vegetation was also estimated. 5.2.2 LANDSCAPE ENVIRONMENTAL FACTORS Landscape indices measured for each golf course included course area, vegetation area, ‘core’ vegetation area (i.e. vegetation buffered by at least 15m of edge habitat), the area of water, patch size, patch perimeter/area ratio (i.e. patch shape) and the area of vegetation within 20m of each patch (Fig. 5.1). Fifteen metres was essentially an arbitrary distance used to distinguish between edge and core habitat and is unlikely to be ecologically meaningful for all vertebrates. It was however a meaningful distance within the context of golf course environments, given the fact that many golf courses are dominated by narrow patches (<15m wide). Aquatic landscape measurements were recorded including the number and area of ephemeral ponds, permanent ponds and streams and the area of vegetation within 100m of each waterbody.
5.2.3 REGIONAL ENVIRONMENTAL FACTORS Regional terrestrial characteristics were measured including the proportion of vegetation cover in an adjacent 200m, 500m, 2km and 5km radius of the course boundary, the proportion of built land cover within a1km radius of each course and the area of vegetation connected within a 200m, 500m, 2km and 5km radius of each course boundary (Fig. 5.1). Regional aquatic characteristics were measured including the area of water in a 200m and 500m radius and the number of stream tributaries connected to the course within a 200m radius.
59
Golf Course Area Patch Area
Core Area On-Course Connectivity
Area of Adjacent Vegetation Area of Connected Vegetation
200m 200m 500m 500m 2000m 2000m 5000m 5000m
Figure 5.1 Schematic diagrams showing the methods used to measure landscape and regional indices from rectified aerial photographs in MapInfo GIS.
60 5.3 RESULTS 5.3.1 TERRESTRIAL HABITATS Golf courses ranged in size from 21ha (G1) to 187ha (G19). Most were between 40- 70ha (Fig. 5.2A). The area of vegetation retained in rough and out-of-play areas varied among golf courses. Most courses devoted only 20-50% of their land area to native vegetation. One golf course (G20), devoted less than 10% of its area to native vegetation, retaining only 5ha. Most golf courses retained between 10-30ha of native vegetation (Figure 5.2A, 5.8). However, the two largest golf courses (G19 and G18) retained 122ha and 98ha respectively (Fig. 5.2A). Very few golf courses retained any substantial area of core vegetation. Much of the habitat found on golf courses was within 15m of cleared land and is therefore likely to be edge-affected. Two golf courses (G19 and G18) were a substantial source of core habitat, supporting 81ha and 71ha respectively (Fig. 5.2A). Seven golf courses (G2, G3, G11, G13, G14, G16 and G17) retained 1.5-4ha of core habitat. The remaining 11 golf courses retained less than 1ha of core vegetation and were therefore dominated by edge habitat. Vegetation patches on most golf courses were smaller than 1ha (Fig. 5.2B). Three golf courses (G10, G6 and G8) did not contain patches larger than 1ha. Another four courses (G9, G1, G12 and G4) held only one patch larger than 1ha. Intermediate- sized patches (2-10ha) were retained on eleven golf courses and three larger patches (>40ha) were retained on the two largest golf courses (Fig. 5.2B). Areas of core habitat on six golf courses (G2, G3, G13, G14, G18, G19) accounted for more than 20% of all core habitat in a 500m radius and therefore represented a substantial local source of habitat (Fig. 5.2.C). The area of native vegetation retained on golf courses was clearly limited by course size. Small golf courses retained fewer large vegetation patches (Fig. 5.3A) than larger golf courses and the total area of native terrestrial vegetation retained on golf courses increased with course size (Fig. 5.3B) While larger golf courses generally devoted a greater proportion of their land area to native vegetation, there was some variation in the relationship (Fig. 5.3C). Courses with significantly less vegetation than expected (based on their size), included G20, G8, G7, G12, G10 and G15. Courses with that retained proportionally more vegetation (for their size) included G1, G3, G13, G11 and G14 (Fig. 5.3C). There was substantial within-site variation in terrestrial habitat complexity. Indices including FHD, CGF, % Natural Ground Cover, tree density, dead tree density and hollow density spanned a comparable range on most sites (Fig. 5.4). All courses,
61 (even those that retain relatively large areas of complex vegetation) had patches of terrestrial habitat with low structural complexity. Likewise, courses with little vegetation cover had high structural complexity in at least a number of patches. Golf courses that retained larger areas of native vegetation however, tended to have greater mean structural complexity, particularly in terms of canopy cover (CGF), foliage height diversity (FHD), the % of natural ground cover, the % of native grass cover and a reduced % of turfgrass cover (Fig. 5.4). There was considerable variation in the mean density of trees, dead trees and hollows among sites (Fig. 5.4). There was a relationship between patch width and habitat complexity. Patches wider than 50m invariably had high structural complexity (Fig. 5.5). Most patches were however less than 50m wide and had highly variable structural complexity. 5.3.2 AQUATIC HABITATS Waterbodies were prominent on most golf courses, accounting for between 2- 6ha of the available land area. Permanent ponds were the dominant waterbody type, accounting for between 30-80% of all waterbodies and 30-95% of the area of water on each course (Fig. 5.6). Streams ran through all but one of the surveyed golf courses (G10) and accounted for a substantial proportion of the local water-surface area on six golf courses (G19, G18, G3, G13, G1 and G7). Ephemeral ponds were rare, occurring on only five of the twenty courses surveyed (G19, G14, G2, G5 and G9) and representing only a relatively small proportion of the total water-surface area on all but two courses (G19 and G14; Fig. 5.6A). Many waterbodies were artificial, with steep hardened banks (made of concrete or rock). Bank edges were in many cases fringed with manicured turfgrass. Aquatic vegetation complexity varied among golf courses (Fig. 5.6B). Courses with low levels of aquatic vegetation complexity included (G11, G13, G4 and G6). In general, golf courses retained between 10-30% vegetation cover in waterbodies. All courses except G13, G1, G4 and G6 retained substantial areas of floating vegetation (mostly in the form of lilies). Courses with relatively little emergent vegetation included (G11, G13, G12, G15 and G6). The proportion of submerged vegetation varied both among golf courses and temporally within individual golf courses. Many courses had periodic increases in submerged vegetation cover (largely due to the prolific growth of algal weeds such as Hydrilla verticilata). This represents a management concern for many golf courses. Hydrilla verticilata is physically removed from many golf courses in SEQ on a periodic basis.
62 A) 200 Water 180 Core Habitat Edge Habitat 160 Turfgrass
) 140 Ha (
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0 0 20 40 60 80 100 120 140 160 180 200 Course Size (Ha) Figure 5.3 Relationship between golf course size and A) patch size, B) total area of vegetation on the course and C) proportion of the course devoted to native vegetation. 64 1.4 120
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30 0.2 20 Foliage Height Diversity Height Foliage
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Figure 5.4. Mean structural complexity of local habitats at each site. Points = means, whiskers = 95%CI. Sites are ordered by total area of vegetation from highest (left) to lowest (right).
65 1.8 0.9
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Figure 5.5 Plots showing the relationship between patch width and the structural complexity of local habitats
66 12 Permanent Ponds Ephemeral Ponds
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Mean % of Bank Frin 0 G19 G18 G3 G14 G17 G2 G11 G13 G16 G1 G12 G4 G5 G20 G9 G7 G8 G15 G6 G10 Site Figure 5.6 Aquatic environmental site variables A) Area of water on each course, B) Mean proportion of vegetation cover, C) Mean proportion of water-body bank fringed by reeds (Columns = Means, Bars = 95% CI, Open circles = outliers) 67 GOLF COURSE BOUNDARY G1
CORE HABITAT
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AQUATICAQUATIC SURVEYS SURVEYS 0 150m Metres
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Figure 5.7 Schematic maps of golf courses G1-G5.
68 GOLF COURSE BOUNDARY G6
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Figure 5.8 Schematic maps of golf courses G6-G10.
69 GOLF COURSE BOUNDARY G11
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Figure 5.9 Schematic maps of golf courses G11-G15.
70 GOLF COURSE BOUNDARY G16
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Figure 5.10 Schematic maps of golf courses G16-G20.
71 5.4 DISCUSSION Golf courses varied in the size and complexity of terrestrial and aquatic habitats. Few golf courses (<5) retained more than 5ha of core terrestrial habitat and only two retained more than 60ha. Most courses retained less than 1ha of core vegetation. Habitat patches were generally small (<1ha) and edge-affected. Golf courses in SEQ are therefore not a substantial source of core eucalypt habitat. As has been found in other studies (Brennan, 1992; Porter et al., 2005), the area of vegetation retained on golf courses increased with golf course size. There was however, some variation in this relationship, with individual courses having substantial room to accommodate local habitat restoration. All golf courses retained areas of over-storey vegetation. There was however, substantial variation in the complexity of mid-level and understorey habitats. While all golf courses had areas of mowed understorey close to playing surfaces, some also cleared understorey vegetation in areas far from playing areas. This was reflected in mean levels of foliage height diversity, understorey height diversity, ground and turfgrass cover and woody debris. This structural simplification of habitats is likely to reduce the diversity of feeding, nesting, breeding and sheltering opportunities and expose wildlife to increased predation, competition and environmental stress. There was little variation in waterbody type with permanent ponds accounting for the majority of wetland habitats on golf courses. Ephemeral and semi-permanent waterbodies were rare. Aquatic wildlife can be influenced by waterbody morphology and hydroperiod. Many wading birds feed only in shallow water (Powell, 1987) and many frog species are specifically adapted to certain waterbody types (Hero et al., 1998). The homogeneity of water-bodies on golf courses may limit species diversity. Variation in the size and complexity of habitats on golf courses reflects the lack of industry regulation governing golf course designs. Few golf developments in SEQ (<1%) have been subjected to EIA scrutiny (Warnken et al., 2001). There has been no obligation to retain native vegetation in new golf developments. Golf courses that have inadvertently retained vegetation face intense development pressure. It is important to determine the impact that the size and complexity of local habitats may have on biodiversity. This will determine the ecological benefits that can be achieved by increasing the area of vegetation retained in future golf developments and by restoring vegetation on existing courses. This potential will be considered in the next chapter.
72 CHAPTER 6 FACTORS INFLUENCING BIODIVERSITY ON GOLF COURSES
6.1 INTRODUCTION A growing number of studies have identified conservation value in small remnants retained in urban and agricultural landscapes (Arnold and Weeldenburg, 1990; Grover and Slater, 1994; Zuidema et al., 1996; Schwartz and van Mantgem, 1997; Sewell and Catterall, 1998; Semlitsch and Bodie, 1998; Fischer and Lindenmayer, 2002a, 2002b). While small remnants have theoretical potential, it is important to define their ecological limitations and ensure small-scale conservation efforts have realistic goals (Linehan and Gross, 1998; Ehrenfeld, 2000). Landscape ecology theory suggests that the ecological value of a given land type can be enhanced by increasing the size, heterogeneity and structural complexity of local habitats (Forman and Godron, 1986; Collinge, 1996). Some studies have however proposed the existence of minimum habitat size-thresholds, below which threatened species diversity is universally low (Howe et al., 1981; Loney and Hobbs, 1991; Edenius and Sjöberg, 1991; Sewell and Catterall, 1998). Efforts to restore or retain habitat on small urban land types may therefore have limited ecological benefit. Researchers have demonstrated that some golf courses can provide refuge to regionally threatened birds (Terman, 1997; Blair, 2001; Merola-Zwartjes and DeLong, 2005; Chapter 2) reptiles, mammals and amphibians (Chapter 2). It is important to determine if the ecological value of other golf courses can be significantly enhanced through improved golf course design and management practices. Golf industry manuals (AGCSA, 1998; AGU, 1998) suggest that wildlife diversity on golf courses can be enhanced by increasing the size and structural complexity of on-course vegetation. These principles have however been adapted from landscape ecology studies conducted at much larger scales. Golf courses are relatively small (60ha on average) and may fall below the critical size thresholds that have been proposed in some landscape ecology studies (eg. Howe et al., 1981; Loney and Hobbs, 1991; Edenius and Sjöberg, 1991). If this is the case, efforts to restore or retain habitat on golf courses may have little more than a cosmetic effect. At the time of this research, ecologists were investigating factors influencing bird diversity on golf courses in the USA (Jones et al., 2005; LeClerc and Cristol, 2005;
73 Merola-Zwartjes and DeLong, 2005; Porter et al., 2005) with varying results. While most found bird species richness increased with the area of forest (Jones et al., 2005; LeClerc and Cristol, 2005) or riparian vegetation (Merola-Zwartjes and DeLong, 2005), others found that on-site management actions had negligible impact and that bird diversity was predominantly determined by the environment surrounding golf courses (Porter et al., 2005). There is currently no information on the environmental factors influencing bird diversity on Australian golf courses or the diversity of reptiles, mammals and amphibians on golf courses anywhere in the world. Fauna surveys conducted on suburban golf courses in SEQ (Chapter 2), present an opportunity to investigate the impact that local habitat size and quality have on the diversity of a range of vertebrates. While a number of golf courses in SEQ have localised conservation value, providing habitat for regionally threatened birds, reptiles, mammals and amphibians, the vast majority support only common urban-adapted fauna and thereby fail to realise that potential. By determining the extent to which the observed variations in biodiversity are attributed to differences in local habitat size and complexity, it would be possible to assess the potential for ecological restoration on existing golf courses and to provide planning guidelines that could be used to enhance the ecological value of future golf course developments. Local biodiversity is however rarely a simple reflection of local habitat conditions. Wildlife assemblages are shaped by a complex interaction of environmental and ecological influences operating at multiple spatial and temporal scales (Forman and Godron, 1986; Kotliar and Wiens, 1990; Dunning et al., 1992). High biodiversity levels observed on some golf courses may therefore be partly explained by external (regional or temporal) influences. This chapter investigates and compares the local, landscape and regional environmental influences affecting biodiversity on golf courses in SEQ. This will provide a greater understanding of the potential ecological benefits of small- scale habitat restoration and conservation efforts on golf courses and other intermediate-sized, semi-natural habitats. 6.2 METHODS Abiotic-biotic relationships were investigated on 20 golf courses in SEQ using both univariate and multivariate analyses. Fauna surveys were conducted at ten randomly selected sampling sub-sites on each golf course using methods outlined in Chapter 2. Abiotic data were measured at three scales at each site: 1) local (within patches),
74 2) landscape (within golf courses) 3) regional (adjacent to golf courses) Information on the methods used to measure abiotic variables is outlined in Chapter 5. For each vertebrate group, a number of biological response variables was assessed including the abundance, species richness, the abundance and species richness of ‘threatened’ (suburban avoiding) species and the composition of entire species assemblages.
6.2.1 UNIVARIATE ANALYSIS Multiple regression analyses were used to test for associations between environmental variables (measured at local, landscape and regional scales) and the site abundance and species richness of birds, reptiles, mammals and amphibians and threatened (urban- avoiding) birds, reptiles, mammals and amphibians and the abundance of specific ecological groups of birds, reptiles, mammals and amphibians. Threatened species were previously identified (Chapter 2), as those significantly less abundant in residential areas of SEQ than in remnants of eucalypt vegetation. Prior to regression, correlation analyses were performed among environmental variables to identify those that would be redundant due to collinearity. Among correlated variables (Pearsons correlation coefficient > 0.7), those considered most ecologically meaningful were retained for analysis. An abbreviated list of potentially explanatory environmental variables was obtained for each vertebrate group (Table 6.1). Table 6.1 Final abiotic variables used in multiple regression analysis
SCALE BIRDS MAMMALS REPTILES FROGS
LOCAL % native grass # trees # logs % floating vegetation FHD # hollows CGF % turfgrass cover # dead trees % native grass % turfgrass cover Mean bank angle # hollows % turfgrass cover CGF # logs # dead trees
LANDSCAPE % covered by veg. Mean patch size Area of vegetation Area of vegetation Area of vegetation Course connectivity Mean patch shape No. of ephemeral ponds Course connectivity Area of veg in 100m Area of water Area of water Area of core veg.
REGIONAL % built land/1km % veg in 2km % built land/1km No. of creeks connected % veg in 200m % veg in 200m % of veg in 500m % veg in 5km % veg in 500m % built land/1km % veg in 5km
75 Multiple regression analyses were used to identify environmental variables that were associated with among-site differences in the abundance and species richness of birds, reptiles, mammals and amphibians and the abundance and species richness of threatened birds, reptiles, mammals and amphibians. Variance Inflation (VIF) values were calculated in each regression as a secondary measure by which to test for collinearity problems. Variables with VIF values exceeding 8 were omitted from the regression. Myers (1986) suggests problems of multicollinearity arise if VIF values exceed 10. Straight multiple regression analyses were used to test associations between local vertebrate abundance and species richness and environmental variables. Regressions were first conducted separately within each scale of potential influence (i.e. local, landscape and regional) and then with all scales combined.
6.2.2 MULTIVARIATE ANALYSIS Multivariate analyses were used to identify environmental factors influencing among- site variation in species assemblages. For each vertebrate group, a data matrix was calculated, representing the relative abundance of each species at each site. Species likely to confound the relationship between species assemblages and environmental features were omitted. These included rare species, flocking species and species with low density and wide distribution (eg. birds of prey). The abundance of remaining species were standardised across sites to provide an index of the relative abundance of individual species at each site. Abundance data were skewed and therefore fourth-root transformed prior to analysis Similarity matrices were calculated for each vertebrate group using the Bray- Curtis similarity index (Bray and Curtis, 1957). An abiotic similarity matrix was then calculated from all measured environmental variables using the Normalised Euclidean similarity index (Wolda, 1981). Possible abiotic-biotic relationships were assessed using BIO-ENV in the PRIMER statistical package (Carr, 1996). BIO-ENV is a multivariate correlation analysis that selects the sub-set of abiotic variables that maximises the rank correlation between biotic and abiotic (dis)similarity matrices. The significance of optimum sub-sets was tested using RELATE (Carr, 1996), a monte- carlo randomisation procedure which permutes the site labels from one of the similarity matrices and recalculates the rank abiotic-biotic correlation many times.
76 6.3 RESULTS
6.3.1 BIRDS: The abundance and species richness of all birds and threatened birds were significantly associated with environmental variables measured at all three scales (Table 6.2 and 6.3). Site bird diversity increased with local environmental indices including the mean proportion of native grass cover (all indices), mean foliage height diversity FHD (species richness) and the number of hollows (bird abundance). Bird abundance and species richness increased significantly with the area of vegetation on the course (Table 6.2 and 6.3). Threatened bird abundance also increased with the level of on-course habitat connectivity (Table 6.3). Regional environmental factors associated with the abundance and species richness of birds (and threatened birds) included the proportion of built land within 1km of each course and the proportion of vegetation cover within a 2km radius (Table 6.2 and 6.3). When all scales were combined, bird abundance was associated with the area of vegetation on the course (Partial R = 0.79) and mean CGF (Partial R = -0.48; Table 6.2, Fig. 6.1). Bird species richness was significantly associated with the area of vegetation on the course (Partial R = 0.53), mean foliage height diversity FHD (Partial R = 0.72,) and the proportion of built land within 1km of the course (Partial R = -0.37; Table 6.2, Fig. 6.1). The site abundance of threatened birds was positively associated with the area of vegetation on the course (Partial R = 0.79) and negatively associated with the proportion of built land within a 1km radius (Partial R = -0.57; Table 6.2, Fig. 6.1). Similarly, the local species richness of threatened birds was associated with the area of vegetation on the course (Partial R = 0.62), the proportion of built land within a 1km radius (Partial R = -0.38) and mean foliage height diversity FHD (Partial R = 0.58). Different ecological bird groups responded to different environmental factors (Table 6.4). One factor (the area of vegetation) was positively associated with the site abundance of all threatened bird groups including hawk gleaners (Partial R =0.63), high gleaners (Partial R = 0.68), pounce gleaners (Partial R = 0.67), understorey nesters (Partial R = 0.78), closed mid-storey nesters (Partial R =0.86) and canopy nesters (Partial R = 0.75). Hawk gleaners were also negatively associated with patch shape (Partial R =-0.60), the % of adjacent built land (Partial R = -0.75) and the % of vegetation within a 200m radius (Partial R = -0.57). High gleaners were negatively associated with the % of vegetation in 200m (Partial R = -0.36). Pounce gleaners were positively associated with CGF (Partial R = 0.52), the area of water on the course
77 (Partial R = 0.37), on-course connectivity (Partial R = 0.29) and negatively associated with the % of vegetation in 2km (Partial R = -0.56). Both understorey nesters (Partial R = 0.52) and canopy nesters (Partial R = 0.32) were positively associated with FHD. The abundance of cleared land birds increased with the % of adjacent built land (Partial R = 0.68), the area of water (Partial R = 0.56) and the % of native grass (Partial R = 0.47).
6.3.2. MAMMALS: Mammal diversity was associated with environmental variables at all scales. Species did however vary in the scale at which they respond to the landscape. The diversity of all mammals was generally most closely associated with local environmental characteristics. In contrast, threatened mammal diversity was more closely associated with landscape-level environmental variables (Table 6.2 and 6.3). Local environmental variables associated with mammal diversity included the number of hollows, the mean proportion of native grass cover, the proportion of turfgrass cover, mean CGF and mean tree density (Table 6.2 and 6.3). Landscape variables associated with the abundance and species richness of mammals (and threatened mammals) included mean patch size, on- course connectivity and the area of water on the course (Table 6.2 and 6.3). Total mammal abundance did not closely reflect regional environmental conditions. Regional factors were however closely associated with among-site differences in mammal species richness and the abundance and species richness of threatened mammals. Influential regional factors included the proportion of built land within 1km and the area of vegetation connected within 2km of each course (Table 6.2 and 6.3). When all scales were combined, mammal site abundance was significantly associated with the mean proportion of native grass cover (Partial R = 0.52), the number of hollows (Partial R = 0.52) and mean tree density (Partial R = 0.46; Table 6.2, Fig. 6.2). Mammal species richness was associated with the mean tree density (Partial R = 0.71), the number of hollows (Partial R = 0.37) and the area of vegetation connected to the course within a 2km radius (Partial R = 0.60; Table 6.2, Fig. 6.2). Threatened mammal abundance was associated with mean tree density (Partial R = 0.79), the mean proportion of turfgrass cover (Partial R = -0.57) and mean patch size
(Partial R = 0.85; Table 6.3, Fig. 6.2). Threatened mammal species richness was significantly associated with the mean tree density (Partial R = 0.45) and mean patch size (Partial R = 0.80; Table 6.3, Fig. 6.2). Different mammal groups responded to different environmental characteristics (Table 6.4). Exotic rodents were most abundant on golf courses that had small patch
78 sizes (Patch Size Partial R = -0.41), a high proportion of built land (Partial R = 0.35), a high % of native grass cover (Partial R = 0.57) and a low % of turfgrass cover (% Turfgrass Partial R = -0.41). In contrast the abundance of small native ground mammals (including native rodents and the yellow-footed antechinus) increased on golf courses with a high % of adjacent vegetation cover within 200m (Partial R = 0.62), a high proportion of native grass cover (Partial R = 0.28), a high canopy gap fraction (CGF Partial R = 0.49) and a low % of turfgrass cover (Partial R = -0.32). The site abundance of common arboreal mammals increased with the abundance of hollows (Partial R = 0.77) and the % of adjacent built land (Partial R = 0.59) and decreased with the level of on-course connectivity (Partial R = -0.64), the area of water (Partial R = -0.63) and the % of vegetation cover within 5km (Partial R = -0.53). Golf courses that supported high macropod densities generally had a combination of large mean patch sizes (Partial R = 0.83), high mean tree density (Partial R = 0.60) and a large area of open space with high canopy gap fraction (Partial R = 0.65), a low % of vegetation cover within 5km (Partial R = -0.55), but a high % of vegetation cover within 2km (Partial R = 0.54).
6.3.3. REPTILES: Reptile diversity was not closely associated with environmental variables measured at any one scale. The abundance and species richness of reptiles (and threatened reptiles) was however more closely associated with local environmental variables than with landscape or regional factors (Table 6.2). Total reptile abundance generally increased with the number of logs and decreased with mean canopy gap fraction CGF (Table 6.2). Local environmental variables associated with differences in the abundance and species richness of threatened reptiles included the mean proportion of turfgrass cover and the number of logs (Table 6.3). Landscape factors that were weakly associated with the abundance and species richness of threatened reptiles included the area of vegetation, the area of water on the course and mean patch perimeter/area ratio (i.e. patch shape). Regional factors weakly associated with the abundance and species richness of threatened reptiles and the species richness of all reptiles included the proportion of built land within a 1km radius, the proportion of vegetation cover within 200m and the area of vegetation connected to the course within 200m (Table 6.2 and 6.3). When all scales were combined, among-site differences in reptile abundance were still most closely associated with mean canopy gap fraction CGF (Partial R = -0.72) and the number of logs (Partial R = 0.75; Table 6.2, Fig. 6.3). Reptile species richness was weakly associated with the number of logs (Partial R = 0.35) and the proportion of
79 vegetation cover in 200m (Partial R = 0.59; Table 6.2, Fig. 6.3). Among-site differences in the abundance of threatened reptiles were associated with mean patch shape (Partial
R = -0.78), the number of logs (Partial R = 0.80) and the proportion of vegetation cover within a 200m radius of each course (Partial R = 0.78; Table 6.3, Fig. 6.3). The species richness of threatened reptiles was significantly associated with local and regional environmental factors including the number of logs (Partial R = 0.75), mean canopy gap fraction CGF (Partial R = -0.76) and the proportion of vegetation cover within a 200m radius of the course (Partial R = 0.79; Table 6.3, Fig. 6.3). Different reptile groups responded to different environmental characteristics (Table 6.4). The site abundance of fossorial reptiles decreased with canopy gap fraction (CGF Partial R = -0.64) and the area of vegetation (Partial R = -0.44) and increased with the area of vegetation connected within 200m (Partial R = 0.49). Threatened skinks were significantly more abundant on golf courses with a high % of vegetation cover within 200m (Partial R = 0.70), a large amount of woody debris (Partial R = 0.68) and rounded patches (Patch Shape Partial R = -0.64). The site abundance of common skinks also increased with the abundance of logs (Partial R = 0.76) patch roundness (Patch Shape Partial R = -0.37) and mean canopy gap (Partial R = -0.54). Common Agamids increased in abundance with the % of adjacent built land (Partial R = 0.71), mean tree density (Partial R = 0.47) and the area of water (Partial R = 0.42) and decreased with the % of native grass cover (Partial R = -0.34). Varanids were significantly more abundant on golf courses with a high proportion of vegetation cover within 500m (Partial R = 0.89), a high tree density (Partial R = 0.48) and with a small area of water (Partial R = -0.42) and canopy gap fraction (Partial R = -0.41).
6.3.4. AMPHIBIANS: Amphibian abundance and species richness were associated with environmental variables measured at all three scales. Influential local variables included mean waterbody bank steepness (i.e. bank angle) and the mean proportion of turfgrass cover on waterbody banks (Table 6.2 and 6.3). Landscape variables significantly associated with amphibian site diversity included the number of ephemeral ponds, the area of terrestrial vegetation and the area of vegetation within 100m of each waterbody (Table 6.2 and 6.3). Significant regional environmental factors associated with amphibian abundance and species richness included the number of creeks connected to the course and the proportion of built land within 1km (Table 6.2 and 6.3). When all scales were combined amphibian site abundance was negatively associated with mean bank
80 steepness (Partial R = -0.49) and the mean proportion of turfgrass cover on waterbody banks (Partial R = -0.61) and positively associated with the mean proportion of floating vegetation cover (Partial R = 0.69; Table 6.2, Fig. 6.4). Amphibian species richness decreased with mean bank steepness (Partial R = -0.70) and with the mean proportion of turfgrass cover on waterbody banks (Partial R = -0.78; Table 6.2, Fig. 6.4). Threatened amphibian abundance increased with the area of terrestrial vegetation (Partial R = 0.74) and the number of ephemeral ponds (Partial R = 0.62; Table 6.3, Fig. 6.4). The species richness of threatened amphibians was associated with the area of terrestrial vegetation on the course (Partial R = 0.44), the number of ephemeral ponds (Partial R = 0.57) and the number of creeks connected to each course (Partial R = 0.51; Table 6.3, Fig. 6.4). As with all other vertebrate groups, different amphibian ecological groups responded to different environmental characteristics (Table 6.4). Ephemeral pond breeders were significantly more abundant on golf courses that retained a higher number of ephemeral or semi-permanent waterbodies (Partial R = 0.61) and which were connected to a larger number of streams (Partial R = 0.57). Threatened active foraging amphibians (eg. Litoria nasuta and Litoria latoplamata) were more abundant on golf courses with a high mean % of floating vegetation cover (Partial R = 0.52), with shallow banks (Mean Bank Angle Partial R = -0.67) and a low abundance of permanent ponds (Partial R = -0.47). The site abundance of regionally threatened tree frogs increased with the level of canopy cover (CGF Partial R = -0.45), the number of permanent ponds (Partial R = 0.52) and decreased with the % of turfgrass cover (Partial R = -0.73) and the mean waterbody bank steepness (Partial R = -0.37). Common treefrog abundance was positively associated with the % of floating vegetation cover (Partial R = 0.57) and negatively associated with the % of turfgrass cover (Partial R = - 0.55) and the area of vegetation within 50m of waterbodies (Partial R = -0.46). The exotic cane toad (Bufo marinus) was most abundant on golf courses that have a high % of adjacent built land (Partial R = 0.78), a high % of reed cover adjacent to waterbodies (Partial R = 0.55), a high % of floating vegetation (Partial R = 0.50) and shallow waterbody banks (Bank Angle Partial R = -0.69).
81
Table 6.2 Significant variables in multiple regression between local, landscape and regional environmental variables and the site abundance and species richness of birds, reptiles, mammals and amphibians. R2 and p values represent overall model statistics, values in brackets represent partial R values. Significant models are shown in bold.
All Species Included
Scale Local Landscape Regional Combined Birds Abundance # Hollows (0.46) Area of veg. (0.80) % Built in 1km (-0.73) Area of veg. (0.79) %Native Grass (0.68) R2 = 0.64, p < 0.0001 % veg in 2km (-0.56) % CGF (-0.48) R2 = 0.54, p = 0.001 Con. Veg - 2km (0.34) R2 = 0.72, p < 0.0001 R2 = 0.55, p = 0.005
Spp Richness % Native Grass (0.38) Area of veg. (0.71) % Built in 1km (-0.72) Area of veg. (0.53) Mean FHD (0.63) R2 = 0.51, p = 0.0004 % veg in 2km (-0.51) % Built in 1km (-0.37) R2 = 0.70, p < 0.0001 Con. Veg - 2km (0.55) Mean FHD (0.72) R2 = 0.63, p = 0.001 R2 = 0.82, p < 0.0001
Mammals Abundance # Hollows (0.52) Area of water (-0.44) % Built in 1km (-0.43) # Hollows (0.52) %Native Grass (0.52) Connectivity (0.56) Con. Veg - 2km (0.26) %Native Grass (0.52) # Trees (0.46) Patch Size (0.28) R2 = 0.34, p = 0.03 # Trees (0.46) R2 = 0.59, p = 0.002 R2 = 0.58, p = 0.003 R2 = 0.59, p = 0.002
Spp Richness Mean CGF (0.54) Connectivity (0.67) % Built in 1km (-0.54) # Hollows (0.37) %Turfgrass (-0.70) R2 = 0.45, p = 0.001 Con. Veg. - 2km (0.54) # Trees (0.71) # Trees (0.36) R2 = 0.57, p = 0.001 Con. Veg. -2km (0.60) R2 = 0.74, p < 0.0001 R2 = 0.71, p = 0.0001
Reptiles Abundance Mean CGF (-0.72) Patch size (0.42) % veg in 200m (0.06) Mean CGF (-0.72) # Logs (0.75) Connectivity (-0.44) R2 = 0.003, p = 0.8 # Logs (0.75) R2 = 0.66, p = 0.0001 R2 = 0.21, p = 0.1 R2 = 0.66, p = 0.0001
Spp Richness % Turfgrass (-0.56) Area of water (-0.22) % Built in 1km (-0.25) # Logs (0.35) R2 = 0.32, p = 0.01 Area of veg. (0.20) % veg. in 200m (0.32) % veg in 200m (0.59) Patch size (0.18) R2 = 0.33, p = 0.03 R2 = 0.37, p = 0.02 R2 = 0.28, p = 0.1
Amphibians Abundance Bank Angle (-0.49) Area of veg. (0.74) % Built in 1km (-0.41) Bank Angle (-0.49) % Turfgrass (-0.61) Area veg 100m (-0.39) # Creeks conn. (0.48) % Turfgrass (-0.61) % Floating veg (0.69) R2 = 0.56, p = 0.001 R2 = 0.51, p = 0.002 % Floating veg (0.69) R2 = 0.71, p = 0.0002 R2 = 0.71, p = 0.0002
Spp Richness Bank Angle (-0.73) Area of veg. (0.61) % Built in 1km (-0.52) Bank Angle (-0.70) % Turfgrass (-0.75) # Ephem. Ponds (0.57) # Creeks conn. (0.54) % Turfgrass (-0.78) R2 = 0.73, p < 0.0001 R2 = 0.64, p < 0.0001 R2 = 0.62, p = 0.0002 R2 = 0.75, p < 0.0001
82
Table 6.3 Significant variables from multiple regression between local, landscape and regional environmental variables and the site abundance and species richness of threatened birds, reptiles, mammals and amphibians. R2 and p values represent overall model statistics, values in brackets represent partial R values. Significant models are shown in bold.
Threatened Species Only
Scale Local Landscape Regional Combined Birds Abundance % Tall Grass (0.74) Area of veg. (0.75) % Built in 1km (-0.71) Area of veg. (0.79) R2 = 0.55, p = 0.0002 Connectivity (0.36) % veg in 2km (-0.40) % Built in 1km (-0.57) R2 = 0.77, p < 0.0001 Con. Veg - 2km (0.40) R2 = 0.82, p < 0.0001 R2 = 0.62, p = 0.001
Spp Richness % Native Grass (0.44) Area of veg. (0.77) % Built in 1km (-0.69) Mean FHD (0.58) Mean FHD (0.47) R2 = 0.60, p < 0.0001 R2 = 0.48, p = 0.001 Area of veg. (0.62) R2 = 0.62, p = 0.0003 % Built in 1km (-0.38) R2 = 0.80, p < 0.0001
Mammals Abundance % Tall Grass (0.42) Area of water (-0.38) % Built in 1km (-0.74) % Turfgrass (-0.57) # Trees (0.53) Connectivity (0.64) R2 = 0.54, p = 0.0002 # Trees (0.79) R2 = 0.51, p = 0.002 Patch Size (0.55) Patch Size (0.85) R2 = 0.76, p < 0.0001 R2 = 0.85, p < 0.0001
Spp Richness # Hollows (0.25) Patch Size (0.63) % Built in 1km (-0.71) # Trees (0.45) %Native Grass (0.60) Connectivity (0.41) Con. Veg. - 2km (0.32) Patch Size (0.80) R2 = 0.40, p = 0.01 R2 = 0.72, p < 0.0001 R2 = 0.62, p < 0.0001 R2 = 0.73, p < 0.0001
Reptiles Abundance % Turfgrass (-0.63) Area of water (-0.36) % Built in 1km (-0.42) # Logs (0.80) # Logs (0.52) Area of veg. (0.52) Con. Veg - 200m (0.25) Patch shape (-0.78) R2 = 0.49, p = 0.003 Patch shape (-0.50) R2 = 0.39, p = 0.02 % veg in 200m (0.78) R2 = 0.46, p = 0.02 R2 = 0.80, p < 0.0001
Spp Richness % Turfgrass (-0.65) Area of water (-0.21) % Built in 1km (-0.27) # Logs (0.75) # Logs (0.43) Area of veg. (0.56) % Veg in 200m (0.32) Mean CGF (-0.76) R2 = 0.49, p = 0.003 R2 = 0.34, p = 0.04 R2 = 0.35, p = 0.03 % veg in 200m (0.79) R2 = 0.77, p < 0.0001
Amphibians Abundance Bank Angle (-0.51) Area of veg. (0.74) % Built in 1km (-0.65) Area of veg. (0.74) % Turfgrass (-0.60) # Ephem. Ponds (0.62) R2 = 0.43, p = 0.002 # Ephem. Ponds (0.62) R2 = 0.53, p = 0.002 R2 = 0.75, p < 0.0001 R2 = 0.75, p < 0.0001
Spp Richness Bank Angle (-0.73) Area of veg. (0.61) % Built in 1km (-0.52) Area of veg. (0.44) % Turfgrass (-0.75) # Ephem. Ponds (0.57) # Creeks conn. (0.54) # Ephem. Ponds (0.57) R2 = 0.73, p < 0.0001 R2 = 0.69, p < 0.0001 R2 = 0.62, p = 0.0002 # Creeks conn. (0.51) R2 = 0.75, p < 0.0001
83 Bird Abundance R2 = 0.72 1600 Partial = 0.79 -20 Partial = -0.48 1500 -80 1400 -140 1300
1200 -200
1100 -260 10 20 40 50 100 0.25 0.35 0.45 0.55 0.65 Area of vegetation (Ha) Mean Canopy Gap Fraction Bird Species Richness R2 = 0.82 76 Partial = 0.53 36 Partial = 0.72 Partial = -0.37 2 30 70
24 -4 64 18 -10 58 12
52 6 -16 10 20 40 50 100 0.20.50.81.1-0.1 0.2 0.5 0.8 Area of vegetation (Ha) Foliage Height Diversity % of built land in 1km Abundance of Threatened Birds R2 = 0.82 840 Partial = 0.79 40 Partial = -0.57 800 0 760 -40 720
-80 680
640 -120 10 20 40 50 100 -0.1 0.1 0.3 0.5 0.7 0.9 Area of vegetation (Ha) % of built land in 1km
Species Richness of Threatened Birds R2 = 0.80 26 84 Partial = 0.62 Partial = 0.58 Partial = -0.38 4 20
76 14 -2
68 8 -8
60 2 -14 10 20 40 50 100 0.20.50.81.1-0.1 0.2 0.5 0.8 Area of Vegetation (Ha) Foliage Height Diversity % of built land in 1km
Figure 6.1. Partial plots for univariate multiple regression between bird diversity indices and combined environmental variables. Dotted line = 95% CI
84 Mammal Abundance R2 = 0.59 1.0
0.8 Partial = 0.46 Partial = 0.52 0.6 Partial = 0.52 0.6
0.4 0.2 0.2
0.0 -0.2 -0.2
-0.4 -0.6 -0.6 20 100 180 260 -10 20 50 -10 10 30 50 Mean number of trees Mean % of native grass cover Number of hollows Mammal Species Richness R2 = 0.71 3 9 4.0 Partial = 0.71 Partial = 0.37 Partial = 0.60 2
7 2.5 1
0 1.0 5 -1
-0.5 -2 3 20 80 140 200 260 320 -10103050 56789 Mean number of trees Mean number of hollows % of vegetation in 2km radius Abundance of Threatened Mammals R2 = 0.85 1.6 5.6 Partial = 0.85 1.2 Partial = 0.57 Partial = 0.78
1.2 5.0 0.8
0.8 4.4 0.4
3.8 0.0 0.4
3.2 -0.4 0.0 0.04 0.1 0.25 1.6 -10103050709020 80 140 200 260 320 Mean patch size Mean % of native grass cover Mean number of trees Species Richness of Threatened Mammals R2 = 0.73 12 Partial = 0.80 Partial = 0.45 1.5 10
0.0 8
6 -1.5 0.04 0.1 0.25 1.6 20 80 140 200 260 320 Mean patch size Mean number of trees Figure 6.2 Partial plots for univariate multiple regression between mammal diversity indices and combined environmental variables. Dotted line = 95% CI
85 Reptile Abundance R2 = 0.66 200 -60 Partial = -0.72 160 Partial = 0.75 -100 120 -140 80
-180 40
-220 0 -260 -40 Mean % of tall grass cover -300 -80 0.25 0.35 0.45 0.55 0.65 -20 40 100 160 Canopy Gap Fraction Number of logs Reptile Species Richness R2 = 0.37 6 Partial = 0.35 Partial = 0.59 4 3
2
0 0
-2
-3 -4 -20 20 60 100 140 180 -0.1 0.1 0.3 0.5 Number of logs % of vegetation in 200m Abundance of Threatened Reptiles R2 = 0.77 -4 12 12 Partial = -0.78 Partial = 0.80 Partial = 0.78 -8 8 8
-12 4 4
-16 0 0
-20 -4 -4 0.1 0.2 0.3 0.4 -20 20 60 100 140 180 -0.1 0.1 0.3 0.5 Perimeter/Area Ratio Number of logs % of vegetation in 200m Species Richness of Threatened Reptiles R2 = 0.77 4 -2 6 Partial = 0.75 Partial = -0.76 Partial = 0.79 4 2 -4
2
0 -6 0
-2 -8 -2 -20 20 60 100 140 180 0.25 0.35 0.45 0.55 0.65 -0.1 0.1 0.3 0.5 Number of logs Canopy Gap Fraction % of vegetation in 200m
Figure 6.3 Partial plots for univariate multiple regression between reptile diversity indices and combined environmental variables. Dotted line = 95% CI
86 Amphibian Abundance R2 = 0.71 -0.2 0.2 Partial = -0.49 Partial = -0.61 0.6 Partial = 0.69 -0.4 0.0 0.4
-0.6 -0.2 0.2
-0.8 -0.4 0.0
-1.0 -0.6 -0.2 40 50 60 70 80 -20 20 60 100 -24 10162228 Mean bank angle Mean % of turfgrass cover Mean % of floating vegetation Amphibian Species Richness R2 = 0.75 -3 Partial = -0.70 2 Partial = -0.78 -5
-7 -2
-9 -6 -11
-13 -10 40 50 60 70 80 -20 0 20 40 60 80 100 Mean bank angle Mean % of turfgrass cover Abundance of Threatened Amphibians R2 = 0.75
1.2 Partial = 0.62 7.6 Partial = 0.74
0.8 7.0
0.4 6.4
0.0 5.8
-0.4 5.2 -1 1 3 5 10 20 40 50 100 Number of ephemeral ponds Area of vegetation (Ha) Species Richness of Threatened Amphibians R2 = 0.75
5 Partial = 0.57 4 Partial = 0.51 13 Partial = 0.44
3 2 11
1 0 9
-1 -2 7 0246-1 1 3 5 10 20 40 50 100 Number of ephemeral ponds Number of creeks connected Area of Vegetation (Ha)
Figure 6.4 Partial plots for univariate multiple regression between amphibian diversity indices and combined environmental variables. Dotted line = 95% CI
87 Table 6.4 Significant variables from univariate multiple regression between environmental characteristics measured at local, landscape and regional scales and the abundance of specific ecological groups of birds, reptiles, mammals and amphibians.
BIRD GROUPS Hawk Gleaners 0.74 % Built land in 1km (-0.75), Area of Veg. (0.63), Patch shape (0.60), % Veg in 200m (-0.57) High Gleaners 0.48 Area of Veg (0.68), % Veg in 200m (-0.36) Pounce Gleaners 0.77 Area of Veg (0.67), % Veg in 2km (-0.56), CGF (0.52), Area of Water (0.35), Connectivity (0.29) Undergrowth Nesters 0.77 Area of Veg (0.78), FHD (0.52) Closed Mid-storey Nesters 0.74 Area of Veg. (0.858) Canopy Nesters 0.69 Area of Veg (0.75), FHD (0.32) Open Mid-storey Birds 0.45 CGF (-0.64), Turfgrass cover (0.42), Area of Veg (0.36), FHD (0.18) Cleared Land Birds 0.5 % Built land in 1km (0.68), Area of Water (0.56), Native Grass (0.47)
REPTILE GROUPS Fossorial skinks 0.47 CGF (-0.60), Area of veg connected in 200m (0.49), Area of Veg (-0.44) Threatened Skinks 0.68 % Veg in 200m (0.70), # of Logs (0.68), Patch Shape (-0.64) Common Skinks 0.67 # of Logs (0.76), CGF (-0.54), Patch Shape (-0.37) Snakes 0.74 Area of Veg (0.69), % Built land in 1km (0.58), % Veg in 2km (0.46), Native Grass (0.33) Common Agamids 0.52 % Built land in 1km (0.71), Tree density (0.47), Area of Water (0.42), Native Grass (-0.34) Varanids 0.8 % Veg in 500m (0.89), # Dead trees (0.48), Area of Water (-0.42), CGF (-0.41)
MAMMAL GROUPS Common Rodents 0.57 Native Grass (0.57), CGF (0.52), % Turfgrass cover (-0.41), Patch Size (-0.40), % Built Land in 1km (0.35) Native Rodents 0.67 % Veg in 200m (0.62), CGF (0.49), % Turfgrass cover (-0.32), Native Grass (0.28) Common Arboreal 0.65 # Hollows (0.77), Connectivity (-0.64), Water Area (-0.63), % Built Land in 1km (0.59), % Veg in 5km (-0.53) Threatened Arboreal NS Insufficient Power to test Macropods 0.82 Patch size (0.83), CGF (0.65), Tree density (0.60), % Veg in 5km (-0.55), % veg in 2km (0.54)
AMPHIBIAN GROUPS Ephemeral Pond Breeders 0.62 # Ephemeral Ponds (0.61), # Connected Streams (0.57) Threatened Active Foragers 0.49 Bank Steepness (-0.67), % Floating Veg (0.52), # Permanent Ponds (-0.47) Threatened Tree Frogs 0.68 % Turfgrass cover (-0.73), # Permanent Ponds (0.52), CGF (0.45), Bank Steepness (-0.37) Common Tree Frogs 0.62 % Floating Veg (0.57), % Turfgrass cover (-0.55), Area Veg within 50m of water (-0.46) Cane Toads 0.66 % Built Land in 1km (0.78), Bank Steepness (-0.69), % Reed cover (0.55), % Floating Veg (0.50)
88 6.3.5. MULTIVARIATE Fifty-three percent of among-site variation in bird assemblage composition (BIO-ENV correlation = 0.53, RELATE p = 0.001) was explained by the mean proportion of turfgrass cover (R = 0.48) and the mean foliage height diversity FHD (R = 0.45; Table 6.5). Fifty-six percent of among-site variation in reptile assemblage composition was explained (BIO-ENV correlation = 0.56; RELATE p = 0.001) by the mean proportion of turfgrass cover (R = 0.28), the proportion of built land within 1km (R = 0.39), the area of vegetation on the course (R = 0.19) and the area of vegetation connected within a 200m radius of each course (R = 0.36; Table 6.5). Forty-five percent of among-site variation in mammal assemblage composition was explained (BIO-ENV correlation = 0.45, RELATE p = 0.001) by mean tree density (R = 0.45; Table 6.5). Fifty-eight percent of among-site variation in amphibian assemblage composition was explained by four variables (BIO-ENV correlation = 0.58; RELATE p = 0.0001). These included the mean proportion of turfgrass cover on waterbody banks (R = 0.44), the number of ephemeral ponds (R = 0.28), the area of terrestrial vegetation on the course (R = 0.34) and the proportion of built land within1km of each golf course (R = 0.36; Table 6.5).
Table 6.5 Results of BIO-ENV multivariate correlation between the local assemblage of birds, reptiles, mammals and amphibians and environmental characteristics.
Vertebrate group Significance Variables (Partial R) Scale
Birds R = 0.53 p = 0.001 % Turfgrass cover (0.48) (Local) Mean Foliage Height Diversity (0.45) (Local)
Mammals R = 0.45 p = 0.001 Mean Tree Density (0.45) (Local)
Reptiles R = 0.56 p = 0.001 % Turfgrass cover (0.28) (Local) Area of vegetation (0.19) (Landscape) % Built land in 1km (0.39) (Regional) Area of veg. Connected in 200m (0.36) (Regional)
Amphibians R = 0.58 p = 0.001 % Turfgrass cover (0.44) (Local) Area of vegetation (0.34) (Landscape) # of ephemeral ponds (0.28) (Landscape) % Built land in 1km (0.36) (Regional)
89 6.4 DISCUSSION Biodiversity is rarely a simple reflection of local habitat conditions (Wiens, 1989; Debinski et al., 2001). Wildlife assemblages are typically determined by a complex interaction of environmental and ecological influences operating at a range of spatial and temporal scales (Forman and Godron, 1986; Turner, 1989; Wiens, 1989). Biodiversity levels on golf courses were therefore unlikely to be determined purely by the size and quality of habitats retained in rough and out-of-play areas. The results of this chapter confirmed this.
6.4.1 REGIONAL INFLUENCES Local biodiversity is often influenced by remote environmental factors (van Druff and Rowse, 1986; Allen et al., 1987; Arnold and Weeldenburg, 1990; Hansson, 1997; Gascon et al., 1999; Rottenborn, 1999; Savard et al., 2000; Debinski et al., 2001; Melles et al., 2004; Porter et al., 2005). These determine the relative isolation of local habitats and the nature and intensity of external suppressive pressures (Savard et al., 2000; Debinski et al., 2001). In this study, biodiversity levels on golf courses were associated with regional environmental factors including the area of vegetation adjacent to golf courses, the area of adjacent built land and the number of connecting streams. In this study, regional environmental factors had a greater influence on the local abundance of animals with large home ranges (i.e. birds and mammals) than on animals that have relatively restricted mobility (i.e. reptiles). The local abundance and species richness of mammals and birds generally increased with the proportion of vegetation cover within a 2km radius of golf courses and decreased with the proportion of built land within a 1km radius. In contrast, the local abundance and species richness of threatened reptiles were only weakly associated with the proportion of vegetation cover within a 200m radius of golf courses. Positive associations with the area of adjacent vegetation may represent an underlying dependence on external resources and reduced insularisation, that could increase local productivity, survivorship and recolonisation rates. The negative association with adjacent built land may reflect the suppression of local productivity and survivorship by noise, disturbance and other anthropogenic effects, as well as the isolating effect of encroaching urbanisation, which would limit resource access, gene exchange and the potential for rescue effects (Robinson and Wilcove, 1994; Knight and Gutzwiller, 1995; Chace and Walsh, 2004). Amphibian species richness also appeared to be influenced by isolation effects. Golf courses that
90 were connected to a number of streams tended to have higher species richness than those that were isolated from external watercourses. Regional factors also partly determined the extent to which golf courses were colonised by exotic and urban-adapted species. Exotic rodents and cane toads (Bufo marinus) and urban-adapted reptiles (eg. the eastern water dragon, Physignathus lesueurii) and common arboreal mammals were more abundant on golf courses that were surrounded by built land. Local biodiversity is partly determined by the area of surrounding vegetation and built land. The extent to which golf courses can provide habitat for threatened vertebrate wildlife will therefore be limited to some extent by the nature of adjacent land types. Similar results were obtained by Porter et al., (2005) when investigating the cause of variations in bird diversity among golf courses in Ohio, USA. Unlike that study however, regional factors were not the sole determinant of biodiversity on golf courses in SEQ. Variations in the size, configuration and structural complexity of local habitats also had an effect on local vertebrate diversity.
6.4.2 LANDSCAPE INFLUENCES Regionally threatened birds, mammals, amphibians and to a lesser extent reptiles, responded to variations in the area of vegetation retained on golf courses. This was an important result. Many studies have observed species-area relationships, whereby species richness increases proportionately with vegetation patch size (Moore and Hooper, 1975; Rafe et al., 1985; Tilghman, 1987; Loyn, 1987; Kerley et al., 1996; Catterall et al., 1997; Major et al., 2001; Vanhinsbergh et al., 2001; Drinnan, 2005). The fact that this association persists at smaller scales (i.e. on golf courses) has implications for small-scale habitat conservation. A growing number of studies have suggested the existence of minimum patch- size thresholds below which the diversity of threatened vertebrates is universally low (Howe et al., 1981; Loney and Hobbs, 1991; Edenius and Sjöberg, 1991; Sewell and Catterall, 1998). This would mean that many small-scale conservation efforts would have limited ecological benefit for threatened vertebrates. While threatened bird, reptile, mammal and amphibian species richness declined with reducing remnant size, the results nevertheless indicate that some golf courses retain a sufficient area of vegetation to be utilised by a range of threatened species. These courses may perform a positive conservation role for regionally threatened vertebrates. While these patches may be unable to support independently viable populations, they may have the capacity to act as part of an interconnected network of smaller urban habitat remnants.
91 The association between local amphibian diversity and the area of terrestrial vegetation was also important, highlighting amphibian dependence on habitats beyond the breeding site. Amphibian conservation often focuses too heavily on the breeding site (Semlitsch, 1998). Many post-metamorphic amphibians are however, also dependent on adjacent terrestrial sheltering and feeding sites (Jameson, 1956; Dole and Durant, 1974; Harris, 1975; Beshkov and Jameson, 1980; Hodgkison and Hero, 2002). Amphibian diversity is for this reason, often influenced by terrestrial habitat characteristics such as woodlot area (Hecnar and M’Closkey, 1997; Knutson et al., 1999; Hazell et al., 2000; Hazell, 2004; Herrmann et al., 2005), proximity to woodland (Marsh and Pearman, 1997; Laan and Verboom, 1999; Ficetola and De Bernardi, 2004) and terrestrial habitat complexity (Pough et al., 1987; Dupuis et al., 1995; Mitchell et al., 1997; Maisonneuve and Rioux, 2001). Efforts to enhance amphibian diversity on golf courses must therefore not ignore the ecological role played by terrestrial habitats. Amphibian diversity also often increases with local waterbody diversity (Beebee, 1997; Hero et al., 1998; 2001; Snodgrass et al., 1999; Hazell et al., 2003; Weyrauch and Grubb, 2004; Rubbo and Kiesecker, 2004). Many amphibians only breed in certain waterbody types, having evolved specific strategies to overcome the inherent threats (i.e. predation and environmental stress) that are unique to those waterbodies (Wiggins et al., 1980; Wellborn et al., 1996; Schnieder and Frost, 1996; Hero et al., 1998; 2001; Weyrauch and Grubb, 2004). Waterbody diversity was generally low on golf courses. Few retained semi-permanent or ephemeral waterbodies. Most golf courses therefore accommodate only part of the local amphibian assemblage, failing to retain ephemeral and semi-permanent pond-breeders, many of which are regionally threatened by urbanisation. Similar results have been observed in urban (Rubbo and Kiesecker, 2004) and agricultural (Hazell et al., 2003) landscapes where landscape changes tend to increase waterbody permanence. Golf courses must retain a diversity of waterbody types if they are to support regionally threatened amphibians. Reptile abundance and species richness was associated with mean patch shape (i.e. perimeter/area ratio). Courses with predominantly narrow, linear vegetation patches tended to have relatively low reptile diversity. Interestingly, patch shape affected the abundance of both common and regionally threatened skinks. The association could be partly attributed to the influence of edge-effects. Patch shape partly determines the extent to which edge-effects penetrate patches (Diamond, 1975; Forman and Godron, 1986; Schonewald-Cox and Bayless, 1986; Collinge, 1996) and
92 reptiles have been found to be particularly sensitive to edge-effects (Gambold and Woinarski, 1993; Demaynadier and Hunter, 1998; Sartorius et al., 1999). On golf courses, edge-effects are likely to be exacerbated by course management practices, in which there is a tendency to clear the understorey of narrow patches that lie close to fairways. This would reduce the diversity of available resources and increase exposure to predation and adverse environmental conditions. The relationship between reptile species richness and patch shape may therefore have more to do with understorey maintenance than patch shape per se. We examined whether the spatial arrangement of habitats on golf courses is important, or if high biodiversity can be maintained simply by retaining a sufficient area of habitat. Habitat configuration and connectivity is often a critical determinant of local biodiversity in fragmented landscapes (Verboom and Van Apeldoorn, 1990), as this determines the capacity for wildlife to continue normal foraging and dispersal movements and to provide opportunities for periodic recolonisation, gene exchange and ‘rescue effects’(Brown and Kodric-Brown, 1977; Fahrig and Merriam, 1985; Stamps et al., 1987). Small-scale variations in habitat connectivity (such as those induced on golf courses by variations in course design) may however have limited influence on biodiversity, particularly if wildlife can move freely across fairways and other playing surfaces. The results indicate that while on-course habitat connectivity is unlikely to be an independently critical determinant of local biodiversity, it did improve the association between landscape conditions and the local abundance and species richness of threatened birds and mammals. Golf course designs that maximise local habitat connectivity are therefore likely to benefit threatened birds and mammals.
6.4.3 LOCAL PATCH SCALE INFLUENCES Biodiversity also closely reflected differences in the structural complexity of local habitats. Different vertebrate groups were associated with different structural features. Bird abundance and species richness and the abundance of understorey nesters and canopy nesters increased with Foliage Height Diversity (FHD) and the mean proportion of native grass cover. Other bird studies have observed associations with FHD and with understorey complexity (MacArthur and MacArthur, 1961; Karr and Roth, 1971; Wilson, 1974; Lancaster and Rees, 1979; Baines et al., 1994; Young and Armstrong, 1995; Chace and Walsh, 2004). Birds tend to partition vertical space, foraging and nesting at different levels of vegetation strata (Lancaster and Rees, 1979). Mid-level and understorey vegetation is often ‘tidied’ on golf courses and in other urban areas.
93 This would inevitably reduce the diversity of foraging and nesting habitats and increase exposure to predation, noise, disturbance and perhaps most significantly, to competition with aggressive nectarivores such as the noisy miner (Manorina melanocephala). Competitive exclusion by this species has been blamed for the local decline of many small insectivorous birds from edge-affected Australian forests (Grey et al., 1997; French et al., 2005). Species that nest, shelter and forage in dense mid-level and understorey habitats are selectively disadvantaged by the removal of understorey vegetation (Barrett et al., 1994; Ford et al., 2001). If golf courses are to retain threatened birds, they must preserve the vertical complexity of vegetation wherever possible. Reptile diversity is typically influenced by the complexity of ground-level structures such as leaf-litter thickness, shrub cover, rockiness, the abundance of woody debris and ground-level heterogeneity (Kitchener et al., 1980; Hadden and Westbrooke, 1996; Smith et al., 1996; MacNally and Brown, 2001; Brown, 2001; Jellinek et al., 2004). In this study, reptile species richness and abundance were positively associated with the abundance of coarse woody debris and negatively associated with the mean proportion of turfgrass cover. The abundance of both common and regionally threatened skinks was also positively associated with the abundance of woody debris. Many reptiles utilise habitats among woody debris or in soil conditions created by rotting wood. These species are likely to be significantly disadvantaged by the lack of feeding and sheltering opportunities provided on golf courses where natural ground habitats have been cleared and replaced with manicured turfgrass. One common reptile species, the grass skink (Lampropholis delicata), was however able to persist in high densities in areas with only simple turfgrass cover. High densities of this species meant that total reptile abundance was unaffected by the mean proportion of turfgrass cover. The abundance and species richness of threatened mammals increased significantly with mean tree density and the proportion of native grass cover. Similar associations with vegetation complexity have been observed in other mammal studies (Lindenmayer et al., 1994; Catling and Burt, 1995; Maisonneuve and Rioux, 2001; Rowston et al., 2002) and may reflect an inherent association between increased foliage volume and food availability (Rowston et al., 2002). Many native rodents are dependent on areas of dense ground cover (eg Rattus fuscipes; Warneke, 1971; Braithwaite et al., 1978; Stewart, 1979; Maitz and Dickman, 2001) and wet sedgeland (eg. Rattus lutreolus; Braithwaite and Gullman, 1978; Braithwaite and Lee, 1979). The
94 number of dead and hollow-bearing trees was also important, influencing the local abundance of arboreal marsupials. Tree hollows are often a critical resource, limiting the local density of hollow-dependent marsupials (Meredith, 1984; Smith and Lindenmayer, 1988; Lindenmayer et al., 1990; Traill and Lill, 1990; Smith and Murray, 2003). In many eucalypt varieties, the number of hollows increases with tree age (Lindenmayer et al., 2000). Mature vegetation and dead hollow-bearing trees must therefore be retained if golf courses are to maintain local populations of hollow- dependent mammals. Amphibian assemblages were affected by local habitat characteristics, primarily the proportion of turfgrass cover on waterbody banks. Amphibian abundance increased with the proportion of floating vegetation cover, largely due to the response of one species, the eastern sedge frog (Litoria fallax). Amphibians also responded to variations in the structure of local waterbodies. Artificial (steep-sided) permanent ponds fringed with manicured turfgrass were common on most golf courses. Golf courses with predominantly artificial ponds generally supported fewer amphibian species than those that retained waterbodies with shallower banks and a reduced level of adjacent turfgrass cover. Waterbodies with steep banks are likely to be accessible to a reduced number of species and those surrounded by turfgrass would have reduced protection from predators and fewer calling sites. Other studies have found amphibian diversity increases with aquatic vegetation complexity (Vos and Chardon, 1998; Hazell et al., 2001; Vallan, 2002) and declines where terrestrial vegetation is removed from areas surrounding waterbodies (Beebee, 1981; Verrell, 1987; Raymond and Hardy, 1991). Efforts to construct more natural waterbodies and increase the structural complexity of emergent, floating and adjacent riparian vegetation are likely to enhance local amphibian diversity. The results of this study have revealed a range of local, landscape and regional abiotic-biotic relationships, demonstrating that landscape ecology principles are also relevant to the ecological management of relatively small urban land types. While some studies have suggested that efforts at small-scale habitat restoration and conservation may be ineffective due to the existence of minimum size-thresholds, this study provides evidence to the contrary. While size thresholds may exist and will inevitably exclude some threatened species from habitats on even the most well-vegetated golf courses, many regionally threatened species have persisted on well-vegetated golf courses in SEQ for a period of several decades.
95 Biodiversity on golf courses can be retained and even enhanced by maximising the size, diversity and structural complexity of terrestrial and aquatic habitats and by maximising connectivity to external aquatic and terrestrial habitats. The results therefore support recent North American studies that have shown that golf course design and management practices can significantly increase bird diversity (Jones et al., 2005; LeClerc and Cristol, 2005; Merola-Zwartjes and DeLong, 2005). While recent studies investigating reproductive success in the eastern bluebird (Sialia sialis) have shown that habitats on golf courses can maintain high reproductive output (LeClerc et al., 2005) or have only slightly reduced productivity (Stanbeck and Seifert, 2005), reproductive success will vary among taxa. More research is required to assess local productivity, threats and the long-term viability of populations on golf courses. This will help to determine the potential risks and benefits associated with connecting potentially sub-optimal habitats on golf courses to more productive habitat networks. Nevertheless, the results demonstrate that golf courses are not too small to retain habitat that is utilised by regionally threatened vertebrates. The ecological value of small-scale habitat conservation and restoration efforts should therefore not be under-estimated.
96 CHAPTER 7 SPECIES-AREA RELATIONSHIPS ON GOLF COURSES
7.1 INTRODUCTION Small habitat remnants can play an important localised conservation role in urban and agricultural landscapes (Franklin, 1993; Zuidema et al., 1996; Fischer and Lindenmayer, 2002a, 2002b). There is however, likely to be a point at which habitats become too small to be of value to threatened vertebrates. Many studies have identified species-area relationships, whereby species richness declines with decreasing remnant size (Moore and Hooper, 1975; Rafe et al., 1985; Tilghman, 1987; Loyn, 1987; Major et al., 2001; Drinnan, 2005). Species-area relationships can be non-linear and some suggest there are distinct minimum patch size-thresholds below which threatened species are lost from the landscape (Loney and Hobbs, 1991; Catterall et al., 1997; Sewell and Catterall, 1998; Drinnan, 2005). Others have found no evidence of a threshold effect (Parker and MacNally, 2002; Lindenmayer et al., 2005) and suggest that species richness simply declines monotonically with decreasing remnant area (Usher, 1987). Information on patterns of small-scale habitat use by threatened vertebrates is required to define the spatial limits for effective conservation action. The previous chapters have demonstrated that small habitats retained on golf courses can be utilised by regionally threatened vertebrates and that the local species richness of threatened vertebrates generally increases with the area of vegetation retained locally. The relationship between species richness and vegetation area has however, not been explicitly demonstrated, given the need to analyse abiotic-biotic relationships at a site level (pooling sub-site abundance), due to low densities of reptiles, mammals and amphibians. This chapter examines species-area relationships on golf courses at a finer level of resolution, by comparing the species richness of birds, reptiles and mammals in different sized remnants (patches) retained in rough and out- of-play areas. This will provide golf course architects, environmental consultants and urban planners with spatial guidelines that could be used to ensure future golf course developments (or restoration efforts) retain remnants that are large enough to be utilised by regionally threatened species.
97 7.2 METHODS
7.2.1 DATA COLLECTION The relative patch density of birds, reptiles and mammals was determined from four surveys (birds), six surveys (reptiles) and two surveys (mammals) using survey techniques described in Chapter 2. Amphibians were not included in this chapter, since amphibians present on golf courses tended to occur only in aquatic sub-sites that were spatially independent of terrestrial vegetation patches. Patch size data was obtained from rectified aerial site photos (Mapview, 2002) using MapInfo GIS.
7.2.2 MINIMUM PATCH OCCUPANCY AND OCCUPANCY RATES For each vertebrate group (reptiles, mammals and birds) a species-accumulation curve was created based on the estimated minimum patch size occupied by individual species (calculated as the mean of the three smallest patches occupied). Estimated minimums were used instead of absolute minimums to account for variation in the behaviour of individual animals and the potential that some observations may have recorded animals during a flight-response (to the observer) and therefore documented abnormal patch occupancies. Patch occupancy rates (% of patches occupied) were calculated for each species over five different patch size intervals (0-0.1ha, 0.1-0.5ha, 0.5-1ha, 1-5ha, 5- 60ha). These were compared with occupancy rates in ten larger stands of continuous eucalypt forest (60-1000ha) within the region.
7.2.3 CHANGE IN THE COMPOSITION OF AVOIDERS, TOLERATORS AND EXPLOITERS Area-dependent changes in assemblage composition were investigated. Species were classified into suburban-avoiders, suburban-tolerators and suburban-exploiters based on their capacity to persist in urban environments in SEQ (Chapter 2). Area-dependent changes in the relative density of suburban-tolerators, exploiters and avoiders were then assessed, comparing their relative composition among the bird, reptile and mammal assemblage found in patch size intervals (0-0.1ha, 0.1-0.5ha, 0.5-1ha, 1-5ha, 5-60ha and 60-1000ha forests). 7.3 RESULTS
7.3.1 MINIMUM PATCH OCCUPANCY
Birds: Species varied substantially in the minimum patch size occupied. Twelve bird species were regularly observed on open land (i.e. fairways and greens) and therefore had no minimum patch requirement for occupancy (Fig. 7.1). These included typical urban-exploiters and tolerators such as the spotted turtle-dove, crested pigeon, galah, noisy miner, feral pigeon, welcome swallow, Australian magpie, magpie-lark, common
98 Brown Cuckoo-dove Grey Crowned Babbler Urban avoider - Fantailed Cuckoo Australian King Parrot Red-browed Firetail Eastern Rosella Urban avoider - Leaden Flycatcher Spotted Pardalote Brush Cuckoo 80 Satin Flycatcher Forest Kingfisher Urban tolerater - Little Wattlebird Urban avoider - Eastern Yellow Robin Urban avoider - White-browed Scrubwren Urban avoider - Restless Flycatcher Weebill Brown Treecreeper Urban avoider - Spangled Drongo Urban avoider - Rufous Fantail Urban avoider - White-throated Treecreeper Mistletoebird Urban avoider – Golden whistler Urban avoider - Grey Fantail Varied Triller Urban avoider - Yellow-faced Honeyeater Urban avoider - Brown Thornbill Chestnut-breasted Mannikin 60 Horsfield's Bronze Cuckoo Urban avoider -Cicadabird Pacific Baza Large-billed Scrubwren Tawny Frogmouth Australian Brush-turkey Urban avoider - Lewin's Honeyeater Urban avoider - Peaceful Dove Urban avoider - Scarlet Honeyeater Urban avoider - Rainbow Bee-eater Urban avoider - White-throated Gerygone Urban tolerator - Eastern Whipbird Pheasant Coucal Urban avoider - Red-backed Wren Urban avoider - Rufous Whistler Azure Kingfisher Urban avoider - Double-barred Finch Urban avoider - Mangrove Gerygone Urban avoider - White-throated Honeyeater Urban tolerator - Variegated Wren 40 Urban avoider - Channel-billed Cuckoo Urban avoider - Collared Kingfisher Golden-headed Cisticola Urban avoider - Grey Shrike-thrush Urban exploiter - Blue-faced Honeyeater Urban tolerator - Silvereye Urban tolerator - Noisy Friarbird Little Friarbird Urban tolerator - Bar-shouldered Dove Urban avoider - Sacred Kingfisher Urban avoider - Superb Blue Wren Urban avoider - Laughing Kookaburra Urban tolerator - Common Koel Cumulative Species Richness Urban avoider - Tree Martin Urban tolerator - Pied Currawong Urban avoider - Striated Pardalote Urban avoider - Dollarbird Urban avoider - Olive-backed Oriole Urban tolerator - Brown Honeyeater Urban tolerator - Sulphur -Crested Cockatoo 20 Urban tolerator - Grey Butcherbird Urban tolerator - Pale-headed Rosella Urban tolerator - Scaly-breasted Lorikeet Urban exploiter - Figbird Urban tolerator - Pied Butcherbird Urban tolerator - Rainbow Lorikeet Urban tolerator - Torresian Crow Urban tolerator - Black-faced Cuckoo-shrike Urban exploiter - Magpie-lark Urban exploiter - Common Myna Urban exploiter- Willy Wagtail Urban exploiter - Noisy Miner Urban exploiter - Magpie Urban exploiter - Crested Pigeon Urban exploiter - Spotted T-Dove Urban tolerator - Galah Urban tolerator - Little Corella Urban exploiter- Welcome Swallow Urban exploiter - Feral Pigeon Urban exploiter - House Sparrow 0 0.01 0.1 1 10 100 Estimated Minimum Patch Size Occupied (Ha)
Figure 7.1 Minimum patch size occupied by individual bird species
99 myna, willy wagtail and little corella. Nine other suburban tolerant birds occupied patches smaller than 0.1ha. Species that are regionally declining due to urbanisation (i.e. suburban-avoiders) occupied minimum patch sizes ranging from 0.2-60ha. There was no distinct threshold in the species-area relationship in birds. Bird species richness declined gradually as patch size decreased, with 42 forest birds disappearing from the landscape between 0.5-5ha. Reptiles: Reptile species richness declined gradually with decreasing patch size, with the majority of suburban-avoiding reptile species disappearing from the landscape between 0.2-5ha (Fig 7.2). Reptiles occupying very small patches (<0.1ha) included suburban tolerant and exploiting species such as the eastern water dragon (Physignathus lesueurii), eastern water skink (Eulamprus quoyii), wall skink (Cryptoblepharus virgatus), grass skink (Lampropholis delicata), bearded dragon (Pogona barbata), Verreaux’s skink (Anomalopus verreauxii) and one suburban avoider, the keelback snake (Tropidonophis mairii; Fig. 7.2). Reptiles that utilised mid- sized patches (0.1-2ha) included the common tree snake (Dendrelaphis punctulata), yellow-faced whip snake (Demansia psammophis), eastern two-lined dragon (Diporiphora australis), scaly-snouted skink (Calyptotis scutirostrum), secretive skink (Lampropholis amicula), lively skink (Carlia vivax), Martin’s skink (Eulamprus martini), robust skink (Ctenotus robustus) and Burton’s legless lizard (Lialis burtonis). Suburban-avoiding reptiles only found in larger patches (>2ha) included the copper- tailed skink (Ctenotus taeniolatus), carpet python (Morelia spilota), short-limbed snake skink (Ophioscincus truncatus), robust rainbow skink (Carlia scmeltzii) and lace monitor (Varanus varius). Mammals: Mammal species richness also declined with decreasing patch size. As with birds and reptiles, there was no distinct threshold in the species-area relationship. Instead, mammal species richness declined gradually, with the majority of threatened mammals disappearing from the landscape between 1-5ha (Fig. 7.3). Exotic pest species including the house mouse (Mus musculus) and brown hare (Lepus capensis) had no lower patch size limit, occurring on open fairways. One suburban-avoider, the eastern grey kangaroo (Macropus giganteus) was also often observed feeding and resting on fairways on a number of golf courses. While this species had no lower patch size limit, it favoured larger patches, with patch occupancy rates for this species, highest in patches >5ha (Append. 9). Suburban-avoiding mammals occupying relatively small patches (<5ha) included the koala (Phascolarctos cinereus), swamp rat (Rattus
100
Urban-avoider Ctenotus taeniolatus
Urban-avoider Carpet python Morelia spilota 20 Urban avoider Blue-tongue lizard Tiliqua scincoides
Urban avoider Short-limbed snake skink Ophioscincus truncatus
Urban avoider Robust rainbow skink Carlia schmeltzii
Urban avoider Lace monitor Varanus varius
Urban avoider Burtons legless lizard Lialis burtonis
Urban tolerator Robust skink Ctenotus robustus
Urban avoider Martin’s skink Eulamprus martini
Urban avoider Lively skink Carlia vivax
Urban avoider Secretive skink Lampropholis amicula
Urban tolerator Scaly-snouted skink Calyptotis scutirostrum
Cumulative Species Richness 10 Urban avoider Eastern two-lined dragon Diporiphora australis
Urban avoider Yellow-faced whip snake Demansia psammophis
Urban avoider Common tree snake Dendrelaphis punctulata
Urban tolerator Verreaux’s skink Anomalopus verreauxii
Urban tolerator Eastern water skink Eulamprus quoyii
Urban tolerator Bearded dragon Pogona barbata
Urban exploiter Grass skink Lampropholis delicata
Urban avoider Keelback snake Tropidonophis mairii
Urban exploiter Wall skink Cryptoblepharus virgatus
Urban tolerator Eastern water dragon Physignathus lesueurii 0 0.1 1 10 100 Estimated Minimum Patch Size Occupied (Ha)
Figure 7.2 Minimum patch size occupied by individual reptile species
101 lutreolus), yellow-footed antechinus (Antechinus flavipes), bush rat (Rattus fuscipes), red-necked wallaby (Macropus rufogriseus) and the squirrel glider (Petaurus norfolcencis). Species only occurring in patches larger than 5ha on golf courses included the pale field rat (Rattus tunneyi) and swamp wallaby (Wallabia bicolor). The greater glider (Petauroides volans) was only found in larger eucalypt forests (60- 1000ha) and was not observed on golf courses.
7.3.2 PATCH OCCUPANCY RATES Many species utilised small patches (<1ha) but had substantially higher occupancy rates in larger patches (>5ha). These included birds such as the sacred kingfisher, dollarbird, olive-backed oriole, bar-shouldered dove, striated pardalote, scarlet honeyeater, cicadabird and rainbow bee-eater, reptiles including the lace monitor (Varanus varius), lively skink (Carlia vivax) and secretive skink (Lampropholis amicula) and mammals such as the bush rat (Rattus fuscipes), eastern grey kangaroo (Macropus giganteus), red-necked wallaby (Macropus rufogriseus), squirrel glider (Petaurus norfolcensis) and yellow-faced antechinus (Antechinus flavipes; Appendix 7-9).
7.3.3 CHANGE IN ASSEMBLAGE COMPOSITION Assemblage composition changed with patch size. Area-dependent changes in the relative abundance of suburban-exploiters/tolerators/avoiders varied among vertebrate groups. Suburban tolerating and exploiting birds dominated the assemblages in smaller patches (<1ha; Fig. 7.4). Suburban exploiting and tolerating reptiles were dominant in all patch size categories including larger forests. And while suburban tolerating and exploiting mammals occurred in relatively low densities in all patch size categories, a relative absence of suburban-avoiding mammals meant that urban-adapted species still dominated the mammal assemblage in smaller remnants. In all vertebrate groups, the composition of urban-avoiders/tolerators/exploiters changed with increasing remnant size. This was principally due to an increase in the relative density of suburban- avoiders, rather than any decline in the density of exploiters or tolerators. Suburban avoiding species only occurred in high relative density in patches >1-5ha (for birds and mammals) and >5ha (for reptiles). Suburban avoiding species of all vertebrate groups occurred at higher densities in medium-sized patches (>5ha on golf courses) than in larger eucalypt forests (60-1000ha).
102
15 Urban avoider Swamp wallaby Wallabia bicolor Urban avoider Pale field rat Rattus tuneyi
Urban avoider Squirrel glider Petaurus norfolcensis
Urban avoider Red-necked wallaby Macropus rufogriseus Urban avoider Bush rat Rattus fuscipes Urban avoider Yellow-footed antechinus 10 Antechinus flavipes Urban avoider Swamp rat Rattus lutreolus Feral urban avoider European fox Vulpes vulpes Cumulative Species Richness Urban avoider Koala Phascolarctos cinereus Feral urban exploiter Black rat Rattus rattus
5 Urban tolerator Common brushtail possum Trichosurus vulpecula Urban tolerator Common ringtail possum Pseudocheirus peregrineus Feral urban exploiter House mouse Mus musculus Feral urban avoider Brown hare Lepus capensis Urban avoider Eastern grey kangaroo Macropus giganteus 0 0.1 1 10 100 Estimated Minimum Minimum Patch SizePatch Occupied Size Occupied (Ha) (Ha)
Figure 7.3 Minimum patch size occupied by individual mammals
103 100 Birds
80
60
40
20 Cumulative Species Richness Species Cumulative
0 0.05 0.5 5 50 500 Patch Size (Ha) 30 Reptiles 25
20
15
10
5 Cumulative Species Richness Species Cumulative
0 0.05 0.5 5 50 500 Patch Size (Ha) 18
16 Mammals
14
12
10
8
6
4
Cumulative Species Richness Species Cumulative 2
0 0.05 0.5 5 50 500 Patch Size (Ha) Figure 7.4 Species-area accumulation curves for birds, reptiles and mammals based on estimated minimum patch sizes occupied. Unbroken black line = all species, dotted black line = urban avoiding species, grey unbroken line = urban-adapted species. 104 Birds Suburban–avoiding birds 120 24 Exploiters 100 Tolerators 20 Avoiders 80 16
60 12
40 8 Species richness 20 4 Mean density /patch
0 0 0-0.1 0.1-0.5 0.5-1 1-5 >5ha FOREST 0-0.1 0.1-0.5 0.5-1 1-5 >5ha FOREST Patch size (Ha) Patch size (Ha) Reptiles Suburban–avoiding reptiles 4 Exploiters Tolerators Avoiders 12 3
2
6
1 Species richness Mean density /patch
0 0 0-0.1 0.1-0.5 0.5-1 1-5 >5ha FOREST 0-0.1 0.1-0.5 0.5-1 1-5 >5ha FOREST Patch size (Ha) Patch size (Ha) Mammals Suburban–avoiding mammals 18 4 Exploiters Tolerators Avoiders 3 12
2
6 1 Species richness Mean density /patch
0 0 0-0.1 0.1-0.5 0.5-1 1-5 >5ha FOREST 0-0.1 0.1-0.5 0.5-1 1-5 >5ha FOREST Patch size (Ha) Patch size (Ha)
Figure 7.5 Change in density of suburban avoiders, tolerators and exploiters and the species richness of suburban avoiders with increasing patch size. Columns = mean values, whiskers = 95% CI
105 7.4 DISCUSSION
7.4.1 THE SPECIES-AREA ACCUMULATION CURVE This chapter has identified the size at which small vegetation patches on golf courses no longer hold any value for regionally threatened vertebrates. The species richness of birds, reptiles and mammals declined progressively with decreasing patch size. There was however, no evidence of a distinct threshold, with threatened species disappearing from the landscape across a broad range of patch sizes. These ranged between 0.5-5ha for birds, 0.2-5ha for reptiles and between 1-5ha for mammals. The species-area relationship observed in birds was consistent with observations made in other Australian bird studies, with threatened species disappearing from the landscape at patches smaller than 5ha (Howe, 1986; Catterall et al., 1987; Sewell and Catterall, 1998; Drinnan, 2005). While there was also no evidence of a higher threshold (>50ha) above which interior-dependent species only begin to occur, as has previously been suggested (Drinnan, 2005), species did continue to appear at larger patch sizes (>60ha). Information on the minimum patch sizes occupied by threatened reptiles and mammals was important. Few studies have investigated area-dependent changes in the species richness of Australian reptiles and mammals. Studies that have investigated species-area responses in reptiles and mammals have produced conflicting results. While a number of studies have found mammal species richness increases with patch size (Bennett, 1987; Goodman and Rakotondravony, 2000), others have not found a species-area effect (Kerley et al., 1996). Similarly, while studies have found that reptile species richness is unaffected by patch size (Kitchener et al., 1980; Dickman, 1987; Burkey, 1995; MacNally and Brown, 2001; Jellinek et al., 2004), others have observed area-dependent changes in reptile species richness (Sarre et al., 1995; Smith et al., 1996; Driscoll, 2004) and in the density of individual species (Kitchener et al., 1980; Sumner et al., 1999; MacNally and Brown, 2001; Lehtinen et al., 2003; Ishwar et al., 2003; Jellinek et al., 2004). This study demonstrates that in the context of suburban golf courses, birds, reptiles and mammals exhibit species-area relationships and indicates that there are broad patch size ranges below which threatened species do not occur.
7.4.2 PATCH OCCUPANCY OF INDIVIDUAL SPECIES The patch size preferences of individual species generally conformed with existing knowledge. Studies suggest that area sensitivity in Australian birds is related to breeding and migrant status. In SEQ, resident forest birds are believed to be more sensitive to reduced fragment size than seasonal migrants (Bentley and Catterall, 1996).
106 Among migrant species, area sensitivity varies seasonally. Migrant bird species that are sensitive to habitat fragmentation in their southern (NSW and VIC) breeding range (Kavanagh et al., 1985; Howe, 1986; MacNally and Bennett, 1997) have been found to occupy considerably smaller remnants in their northern (SEQ) winter habitats (Bentley and Catterall, 1996). Similar seasonal shifts in area sensitivity have been observed in Nearctic birds in Central and North America (Hutto, 1989; Maurer and Heywood, 1993; Böhning-Gaese et al., 1993). In this study, many forest residents were found only in larger remnants (5-60ha on golf courses or 60-1000ha in forests). These included the buff-rumped thornbill, painted button quail, glossy black cockatoo, powerful owl and brown cuckoo-dove. Two forest residents (the grey shrike-thrush and variegated wren) did however occupy relatively small patches (1-5ha). Species that have been classified as forest migrants or habitat generalists (Sewell and Catterall, 1998; Bentley and Catterall, 1998) were generally found in smaller remnants (<5ha). Small habitats on golf courses therefore appear to have a greater capacity to accommodate migrant birds and habitat generalists than forest residents. Larger forests may be required to protect some forest residents. There is little information on the relative sensitivity of reptile species to reduced habitat size. Species that have been observed to decline following reductions in patch size typically include habitat specialists (Kitchener et al., 1980; Sumner et al., 1999; MacNally and Brown, 2001; Ishwar et al., 2003; Jellinek et al., 2004), species that avoid edge and matrix habitats (Sarre et al., 1995; Lehtinen et al., 2003) and ambush predators that have restricted foraging ranges (Webb and Shine, 1998). Patterns of reptile patch occupancy observed on golf courses generally conformed to those generalisations. Reptiles occupying small remnants were typically urban-tolerant species with broad ecological tolerances (eg. the grass skink, Lampropholis delicata, wall skink Cryptoblepharus virgatus and bearded dragon Pogona barbata), species dependent on other (aquatic) habitats (i.e. eastern water dragon Physignathus lesueurii, eastern water skink Eulamprus quoyii, and the keelback Tropidonophis mairii) or relatively active-foragers (the common tree snake Dendrelaphis punctulata and yellow- faced whip snake Demansia psammophis). Larger reptiles (eg. the carpet python Morelia spilota and lace monitor Varanus varius) and species associated with complex habitats (eg. Ctenotus taeniolatus, Ctenotus robustus, Carlia schmeltzii, Carlia vivax, Ophioscincus truncatus and Lialis burtonis) tended to occupy larger remnants.
107 Area-dependency varied among mammal species. Those generally only found in larger remnants included the greater glider (Petauroides volans) and swamp wallaby (Wallabia bicolor). Greater gliders were not found in any of the golf courses surveyed but did occur in larger forests. This species is known to be particularly sensitive to habitat alteration (Kavanagh and Bamkin, 1995). The swamp wallaby was only observed in remnants larger than 2ha. This solitary species is known to be relatively secretive, browsing in dense vegetation during the day and only feeds in open pasture at night (Edwards and Ealey, 1972). Suburban-avoiding mammals that have been found to at least temporarily utilise small remnants (1-10ha) in previous studies include the bush rat Rattus fuscipes (Laurance, 1991; 1994; Crome et al., 1994; Downes et al., 1997; Bentley et al., 2000), squirrel glider Petaurus norfolcensis (van der Ree et al., 2003), koala Phascolarctos cinereus (Pahl et al., 1990) eastern-grey kangaroo Macropus giganteus (Kaufmann, 1974; Hill, 1981), red-necked wallaby Macropus rufogriseus (Johnson, 1987) and yellow-footed antechinus Antechinus flavipes (Bentley et al., 2000; Marchesan and Carthew, 2004). Each of these species occupied relatively small patches (1-5ha) in this study. All species however, had higher occupancy rates in larger remnants. While the eastern grey kangaroo and red-necked wallaby were observed feeding on open fairways during the day, they rarely ventured far from cover. Similar patterns of habitat use have been observed previously for these species (Caughley, 1964; Frith and Calaby, 1969; Frith, 1973; Fox, 1974; Kaufmann, 1974; Johnson, 1987; Bennett, 1987). Koalas and squirrel gliders also demonstrated a capacity to cross fairways, occurring in small isolated remnants at least 50m from larger vegetation patches. The squirrel gliders’ use of local habitats is restricted by their maximum gliding distance (van der Ree et al., 2003) - approximately 75m. Individuals of the closely related sugar glider (Petaurus breviceps) have however been known to walk up to 250m across areas of open pasture (Suckling, 1984). While the bush rat was not found in open areas (i.e. fairways) in this study or in other previous studies (Crome et al., 1994; Downes et al., 1997; Bentley et al., 2000), it has been known to utilise open pastures adjacent to vegetation in north Queensland (Laurance, 1991; 1994).
7.4.3 CHANGE IN ASSEMBLAGE COMPOSITION Inter-specific differences in the capacity to utilise open ground and small remnants leads to a shift in species composition with increasing patch size. On golf courses, smaller patches were dominated by suburban-tolerant and suburban-exploiting species.
108 As patch size increased, the density of suburban tolerators and exploiters remained relatively constant, however the local density of suburban avoiders increased. Although many regionally threatened vertebrates occupied patches smaller than 1ha, most did not occur in any substantial density until patches were between >1-5ha (for birds and mammals) and >5ha (for reptiles). The fact that threatened reptiles generally had larger minimum area requirements (for occupancy) than threatened mammals and birds was counter-intuitive, given their smaller size and potentially restricted home range. The relative absence of threatened reptiles from patches <5ha may however not be a direct response to patch size. Instead, the association may reflect a co-correlated decline in habitat complexity among smaller patches (where the understorey is often cleared on golf courses). Larger patches may therefore simply be the only areas that retain the microhabitats required by threatened reptiles. Co-correlated complexity-dependence has been used to explain area-dependent declines in the density of Egernia spp. (Kitchener et al., 1980; MacNally and Brown, 2001; Jellinek et al., 2004). Alternatively, the relative absence of suburban-avoiding reptiles from small remnants (<5ha) may reflect an added isolation effect. Species-area relationships are driven partly by island biogeography principles in which species richness is determined by local rates of immigration and extinction (MacArthur and Wilson, 1967; Fisher and Lindenmayer, 2002a). On golf courses, fairways and other open areas may represent a greater barrier to animals that have relatively low dispersal abilities (i.e. reptiles) than to more mobile fauna (i.e. birds and mammals). Patch isolation would exacerbate local declines and prevent recolonisation events, thereby compounding the effects of reduced fragment size. In this study, larger reptiles (eg. snakes, large dragons, monitors and large skinks) were occasionally observed crossing fairways. Smaller reptiles (excluding Lampropholis delicata and Cryptoblepharus virgatus) were however never observed on fairways and may therefore be to some extent isolated within habitat remnants. For all vertebrate groups, suburban-avoiding species occurred at higher densities in medium sized remnants (>5ha on golf courses) than in larger eucalypt forests (60- 1000ha). Similar results have been observed in other studies (Loyn, 1987; Barrett et al., 1994; Bentley and Catterall, 1996; Catterall et al., 1998) and may be attributed to a number of causal factors. Some suggest elevated densities in mid-sized patches may be a residual crowding effect, following the fragmentation of regional habitats (Whitcomb et al., 1981; Fahrig, 1991; Debinski and Holt, 2000). Others suggest medium-sized remnants may have intrinsically higher fertility than larger forests, since development
109 tends to target productive land and overlook less fertile areas (Catterall et al., 1997). Mid-size patches may also support fewer meso-predators which could tend to elevate local densities in prey populations. Alternatively, high mid-patch size densities may reflect differential wildlife detectability among small, mid-size and large patches. These could be driven by the differences in the availability of hiding places (higher in large forests) and animal calling behaviour (more active in mid-size remnants where there is often lower mate-pairing success and thus greater calling activity).
7.4.4 MANAGEMENT IMPLICATIONS This chapter provides spatial guidelines that could be used to ensure future golf course developments in SEQ retain habitats that are large enough to be utilised by regionally threatened species. There are however, several reasons for caution in the application of these results. Firstly, it is critical to point out that the minimum patch size data presented represent the minimum area required for occupancy, not for persistence. Many threatened species observed in this study are utilising patches that may be too small to support independently viable populations. Viability in small remnants will be dependent on local productivity and connectivity to other adjacent habitats. Small patches that form part of a larger network of habitats may be able to support threatened species indefinitely, however they are unlikely to have any long-term value if they occur in isolation (Opdam, 1991; Merriam, 1991; Fahrig and Merriam, 1994). The patch size required for occupancy will therefore tend to increase over time, as isolated populations (occurring below the viability threshold) become locally extinct. The species-area relationships presented should not be regarded as minimum acceptable guidelines by which to manage development impacts (Radford et al., 2005). Instead, they represent the points at which multiple species become locally extinct. They represent points of absolute ecological instability, below which natural systems collapse and should therefore be altogether avoided (Radford et al., 2005). The absence of any distinct threshold highlights the dangers of seeking to identify explicit spatial guidelines for ecologically sound development. In this study, there was no clear patch size above which, most threatened species are represented. Threatened species continued to appear as patch size increased beyond 50ha. While many threatened vertebrates occupied patches as small as 1ha, they invariably occurred at higher density in larger patches. While this chapter provides further evidence to support small-scale vertebrate conservation efforts on golf courses and other urban green-space areas, it demonstrates that the success of such efforts will be dependent on the area of remnant
110 native vegetation retained. Protected areas should therefore always be as large as possible. Species-specific research is required to assess the relationship between patch size and reproductive success and to obtain a spatially explicit understanding of source- sink dynamics and the role that small remnants can play in maintaining functioning metapopulations of threatened species within fragmented urban landscapes.
111
112 CHAPTER 8 CHANGES IN THE LOCAL BIRD ASSEMBLAGE FOLLOWING HABITAT LOSS ON A GOLF COURSE
8.1 INTRODUCTION Golf courses in SEQ face intense development pressure. Between 2003-2005, four of the twenty golf courses surveyed in this study proposed or made a significant reduction in land area to accommodate housing development. Other golf courses in SEQ face similar development pressure. Urban planning authorities in SEQ have raised concerns that the development of land areas originally intended to provide open-space and recreation opportunities may have adverse social impacts through lost recreation opportunities and degraded aesthetic value (McInally, 2005). Clearing small habitat remnants on golf courses for housing will also have negative ecological consequences. In 2003, 9.6ha of native vegetation was cleared from golf course (G17) and replaced with a housing estate (Fig. 8.1). This represented a loss of 30% of the original area of native vegetation present at the site. Prior to the development, the course provided habitat that was utilised by regionally threatened birds and mammals. This chapter documents changes in the local bird assemblage following that development. 8.2 METHODS
8.2.1 BIOTIC AND ABIOTIC SURVEYS Golf course (G17) was surveyed twice for birds at three time periods: before, during (post clearing) and approximately one year after the development (post construction). Bird surveys were conducted at ten randomly selected sub-sites on the course using methods described in Chapter 2. The same sub-sites were surveyed each time. All bird surveys were conducted on clear spring mornings, thereby avoiding seasonal or climatic variations in local bird activity. The area of vegetation cleared for development was estimated using a desk-top area measurement tool using rectified aerial images in Mapview (2002). This was only an estimate, based on comparison of aerial photographs and on-site observation.
8.2.2 DATA ANALYSIS Differences in the site abundance and species richness of all birds and the abundance of suburban-avoiding, tolerating and exploiting birds were compared before and after the development using a one-way ANOVA. Change in the local bird assemblage was
113 illustrated graphically using MDS ordination (Belbin, 1996), plotting species composition at the site (before and after the development), relative to the species composition of bird assemblages observed on other golf courses, eucalypt remnants and suburban residential areas within the region. 8.3 RESULTS
8.3.1 CHANGE IN BIODIVERSITY Local bird abundance was substantially higher after the development ( x = 217, SD =
58) than prior to construction ( x = 113.5, SD = 21.6, p = 0.1). Bird species richness was substantially lower after the development ( x = 11.5, SD = 2.1) than it was previously ( x = 18, SD = 1.9, p = 0.07). The local bird assemblage was therefore substantially homogenised by the development. Few species persisted, but those that did occurred in high local densities. Suburban-exploiting birds were substantially more abundant after the development ( x = 123.5) than before ( x = 23. p = 0.06) or during construction ( x = 16.5, p = 0.06). Suburban-avoiding birds were significantly less abundant after the development ( x = 3.5) than they had been before ( x = 17.5, p = 0.02) or during construction ( x = 12, p = 0.06). While there was no difference in the abundance of suburban tolerating birds before or after the development (Fig. 8.2), there was insufficient power to detect any difference (0.07).
8.3.2 SPECIES RESPONSES Two species, the noisy miner (Manorina melanocephala) and Torresian crow (Corvus orru) were substantially more abundant after the development than they had been previously (Figure 8.3). Only 2 of the 18 suburban-avoiding bird species found locally prior to development were found after construction. These included the sacred kingfisher (Todiramphus sanctus) and the dollarbird (Eurystomus orientalis). Suburban-avoiding species not found after the development included the white-throated treecreeper (Cormobates leucophaea), white-throated gerygone (Gerygone olivacea), superb blue wren (Malurus cyaneus), scarlet honeyeater (Myzomela sanguinolenta), striated pardalote (Pardalotus striatus), rainbow bee-eater (Merops ornatus), pheasant coucal (Centropus phasianinus), olive-backed oriole (Oriolus sagittatus), laughing kookaburra (Dacelo noveguineae), grey shrike-thrush (Colluricincla harmonica), grey fantail (Rhipidura fuliginosa), eastern whipbird (Psophodes olivaceus), eastern rosella (Platycercus eximius), channel-billed cuckoo (Scythrops novaehollandiae), Australian king parrot (Alisterus scapularis) and the variegated wren (Malurus assimilis).
114
Pre-development
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Developed Area
Figure 8.1 Schematic map of golf course G17 before and after housing development
115 200 Suburban-Exploiting Birds 160
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40 Pre-development Construction Phase Post-development
30 Suburban-Avoiding Birds
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Figure 8.2 Change in site abundance of suburban-exploiting, tolerating and avoiding birds at golf course G17 before, during and after development. Boxes = mean abundance, whiskers = max-min abundance values from two repeated surveys.
116 200 Before development
160 Urban Urban Urban Exploiters Tolerators Avoiders 120
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160 Urban Urban Urban 120 Exploiters Tolerators Avoiders
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Galah Galah Figbird Figbird Dollarbird Dollarbird Magpie-lark Grey Fantail Noisy Miner Noisy Noisy Miner Magpie-lark Magpie-lark Willy Wagtail Willy Grey Fantail Common Koel Common Willy Wagtail Common Koel Common Myna Common Crested Pigeon Crested Friarbird Noisy Crested Pigeon Pigeon Crested Torresian Crow Eastern Rosella Little Wattlebird Little Eastern Rosella Common Myna Noisy Friarbird Pied Currawong Torresian Crow Variegated Wren Pheasant Coucal Pied Currawong Pheasant Coucal Pied Butcherbird Grey Butcherbird Grey Little Wattlebird Rainbow Lorikeet Rainbow Eastern Whipbird Eastern Variegated Wren Pied Butcherbird Striated Pardelote Sacred Kingfisher Grey Butcherbird Brown Honeyeater Superb BlueWren Sacred Kingfisher Rainbow Lorikeet Rainbow Bee-eater Scarlet Honeyeater Eastern Whipbird Striated Pardalote Grey Shrike-thrush Grey Superb Blue Wren Brown Honeyeater Brown Honeyeater Rainbow Bee-eater Olive-backed Oriole Olive-backed Scarlet Honeyeater Honeyeater Scarlet Spotted Turtle-Dove Spotted Grey Shrike-thrush Pale-headed Rosella Pale-headed Rosella Olive-backed Oriole Spotted Turtle-Dove Black-backed Magpie Bar-shouldered Dove Bar-shouldered Dove Laughing Kookaburra Laughing Black-backed Magpie Australian King Parrot King Australian Channel-billed Cuckoo Channel-billed Laughing Kookaburra Scaly-breasted Lorikeet Channel-billed Cuckoo Australian King Parrot Scaly-breasted Lorikeet White-throated Gerygone White-throated White-throated Gerygone White-throated Black-faced Cuckoo-shrike Sulphur Crested Cockatoo Sulphur-crested Cockatoo White-throated Treecreeper Black-faced Cuckoo-shrike White-throated Treecreeper
Figure 8.3 Changes in the abundance of bird species before, during and after an on-course housing development at golf course G17.
117
Before Development
Suburbia
Golf Courses Axis 2
Eucalypt Forests
Axis 1
After Development
Suburbia
Golf Courses Axis 2
Eucalypt Forests
Axis 1
Figure 8.4 SSH-MDS Ordination plot showing the change in bird assemblage composition on golf course G17 (circled) relative to other golf courses (blue triangles), suburban residential areas (green squares) and eucalypt remnants (red circles). Ellipses represent 90% confidence intervals around each land type.
118 8.3.3 SPECIES COMPOSITION Prior to the development, the bird assemblage on golf course G17 was between that of a well-vegetated suburb and a low quality eucalypt forest (Fig. 8.4). After the development the bird assemblage was substantially homogenised, containing a greater relative abundance of urban-exploiting birds than were found in most residential areas. 8.4 DISCUSSION Urban bird diversity has been found to increase with the area of native vegetation cover (Emlen, 1974; Mills et al., 1989). Any urban development that results in a substantial loss of vegetation cover will therefore have a significant homogenising effect on urban bird assemblages (Blair, 1996). In this process a diverse range of bird species, dependent on specific nesting and feeding sites provided by complex, stratified native vegetation are replaced by a small suite of ecological generalists and urban-adapted bird species, able to tolerate and persist in cleared or structurally simplified habitats (Blair, 1996; McKinney and Lockwood, 1999; Taylor 2001). Given the fact that bird species richness increases with remnant size (Moore and Hooper, 1975; Rafe et al., 1985; Tilghman, 1987; Loyn, 1987), the negative effects of urban land clearing will be most severe when large remnants (supporting diverse bird assemblages) are lost. Small losses of native vegetation (1-2ha) are therefore often presumed to have little ecological impact and are therefore overlooked. The results of this chapter demonstrate that the loss of even relatively small habitat remnants on golf courses can have dramatic consequences for local bird diversity. Prior to the development, golf course G17 supported a relatively diverse bird assemblage including 18 species that are regionally disappearing from SEQ due to urbanisation. While the habitats on golf course G17 were clearly of relatively low conservation value (compared to larger eucalypt remnants found in the region) they would nevertheless represent a refuge for birds within the context of the surrounding residential landscape. After the development, any local refuge value was lost. The site became dominated by urban-adapted bird species including the Torresian crow (Corvus orru), magpie lark (Grallina cyanoleuca), spotted turtledove (Streptopelia chinensis) and willy wagtail (Rhipidura leucophrys). One species, the noisy miner (Manorina melanocephala) experienced an exponential increase in local abundance. This species has been associated with the localised decline of forest birds from fragmented Australian forests (Catterall et al., 1997; Grey et al., 1997; MacNally et al., 2000;
119 Major et al., 2001; French et al., 2005). In this situation, it is however uncertain whether the local decline of forest birds was exacerbated by competition with the noisy miner or if it simply represented a response to local habitat loss. Regardless of the causal mechanisms, it is clear that the housing development (and associated loss of habitat) has had a substantial, homogenising effect on the local bird assemblage. Where this site was previously a suburban refuge for regionally threatened birds, it is now a local strong-hold for noisy miners. The site has therefore gone from having a positive or neutral regional influence on bird diversity to being a potentially negative influence, given the capacity for noisy miners to radiate outwards and suppress threatened bird populations in adjacent remnants. While the loss of small habitats from golf course G17 is unlikely to have substantial regional implications for bird diversity, the ongoing loss of many other similar small remnants from golf courses and open-space land types in the region is likely to have a cumulative negative impact on regional bird diversity. This cumulative effect (associated with the loss of many seemingly insignificant small remnants) is responsible for significant biodiversity declines in urban centres (Theobald et al., 1997). The results of this chapter highlight the danger in underestimating the existing contribution that small habitat remnants can make to local wildlife conservation. Small and mid-sized remnants should be retained on urban land types wherever possible. Clearing vegetation for development on golf courses and other urban open-space areas will have a negative homogenising effect on urban bird assemblages.
120 CHAPTER 9 MANAGEMENT RECOMMENDATIONS
9.1 INTRODUCTION Conservation actions are implemented through a socio-economic rather than ecological process (Freeman, 1999; Luz, 1999). Thus while this study has shown that golf courses in SEQ could perform a positive localised conservation role, the extent to which that potential is realised will depend on land management decisions made at local, landscape and regional levels. Information on the social, economic and logistical factors affecting golf course design and management is required to understand and implement the changes necessary to realise the golf industries conservation potential. This chapter outlines management recommendations that could enhance the ecological value of new and existing golf courses and discusses social and logistical factors that might restrict ecological enhancement at each scale of management and which may therefore need to be overcome. 9.2 LOCAL SCALE ACTIONS:
IMPLICATIONS FOR GOLF COURSE SUPERINTENDENTS The capacity for golf courses to support regionally threatened vertebrates is partly dependent on local habitat complexity. Habitat complexity affects local biodiversity by increasing the abundance and diversity of feeding, nesting, sheltering and breeding opportunities, and increasing protection from predators and aggressive urban-adapted competitors (Forman and Godron, 1986; Collinge, 1996). In areas where the ecological requirements of regionally threatened species are met, these species are likely to have a greater capacity to exclude opportunistic, but competitively inferior exotic species such as the house mouse (Mus musculus; Haering and Fox, 1993). On golf courses, the structural complexity of local habitats is determined by two stages of land management. Initially, habitat complexity is determined by golf course designs, and will vary depending on the extent to which complex vegetation is retained in rough and out-of-play areas. Over time, local habitat complexity is more heavily influenced by the ongoing management actions of golf course superintendents. In this study, habitat complexity was generally low on most golf courses. In many cases, mid- level, understorey and ground-level habitats were structurally simplified (Chapter 5). In some areas, this is a functional requirement, necessary to ensure course playability and to
121 reduce ball loss. On most golf courses however, there were areas (further from playing areas), where habitat complexity has been reduced for aesthetic reasons. Reduced habitat complexity is a critical factor, limiting the conservation value of suburban golf courses. Management actions that could be used by golf course superintendents to improve the conservation value of golf courses are outlined below. Biodiversity on golf courses should increase with: • foliage height diversity (birds) • mean tree density (mammals) • the number of hollow-bearing trees (birds – lorikeets, arboreal mammals) • the proportion of native grass cover (birds – finches, mammals – native rodents) • the number of logs and vegetative debris (reptiles) • the complexity of aquatic and adjacent riparian vegetation (amphibians) and decrease with: • the proportion of turfgrass cover (reptiles, mammals, amphibians) • waterbody bank steepness (amphibians)
9.2.1 INCREASE VERTICAL COMPLEXITY OF VEGETATION Threatened birds require vegetation with complex vertical strata (Table 6.2, 6.3, Fig. 6.1). Birds partition vertical space, foraging, nesting and sheltering at different vegetation heights (Barrett et al., 1994; Ford et al., 2001). Golf courses with predominantly cleared understorey had negligible value for threatened birds and instead supported only common urban-adapted species. Golf courses can increase their capacity to support regionally threatened birds by retaining or restoring complex ground level, understorey, mid-level and canopy vegetation. Even small complex remnants (particularly those connected to larger patches) can be utilised by urban-avoiding birds.
9.2.2 INCREASE MEAN TREE DENSITY Many golf courses have sparse parkland style vegetation with cleared understorey and low tree density. Courses with higher mean tree densities (approximately >200 trees ha-1) supported a higher abundance and species richness of threatened mammals (Fig. 6.2). Areas with elevated tree density are likely to have a greater abundance and diversity of food, foraging sites and nesting sites and offer greater protection from noise, disturbance, predation and competition (Maisonneuve and Rioux, 2001; Rowston et al., 2002). Tree density should therefore be maximised in areas away from playing areas, either by retaining existing native vegetation or through replanting schemes.
122
Figure 9.1 Increasing the vertical complexity of native vegetation will
increase bird, reptile and mammal diversity.
Figure 9.2 Increasing tree density will increase local mammal diversity.
123 9.2.3 RETAIN DEAD TREES AND HOLLOW-BEARING TREES Hollow-bearing trees are a critical factor, limiting the local abundance of hollow-dwelling mammals and birds (Table 6.2, Fig. 6.2; Meredith, 1984; Lindenmayer et al., 1990; Traill and Lill, 1990). Local hollow-dependent species include the common brushtail possum (Trichosurus vulpecula), squirrel glider (Petaurus norfolcencis), greater glider (Petauroides volans), lace monitor (Varanus varius) and birds including lorikeets, parrots, rosellas, cockatoos, tree-creepers, kingfishers, rollers and owls. Local bird and mammal diversity can therefore be increased by retaining dead trees with hollows. Golf courses that have few natural hollow-bearing trees may have some capacity to accommodate hollow-dwelling fauna using nest-boxes. These are more likely to be used by regionally threatened species if they are placed in areas of closed forest. Nesting boxes placed in open areas (on fairways) are only likely to attract urban-adapted species such as the common brushtail possum (Trichosurus vulpecula), rainbow lorikeet (Trichoglossus haematodus) or common myna (Acridotheres tristis).
9.2.4 INCREASE THE PROPORTION OF NATIVE GRASS COVER Golf courses that retain areas of native grass support a higher abundance and species richness of threatened birds and mammals (Table 6.2, 6.3, Fig. 6.2). Native grass provides essential nesting sites, protection from predators, food for granivores and rich feeding sites for insectivores (Ryan, 2000; Cole and Lunt, 2005). Species likely to benefit from the retention or restoration of native grass include small native ground mammals such as the bush rat (Rattus fuscipes), swamp rat (Rattus lutreolus), pale field rat (Rattus tunneyi) and yellow footed antechinus (Antechinus flavipes) and a range of finches and other small granivorous birds.
9.2.5 RETAIN LOGS, FALLEN WOOD AND ROTTING WOODY DEBRIS Fallen wood is often removed from the understorey of rough and out-of-play areas on golf courses. Coarse woody debris provides valuable microhabitat complexity (shelter and feeding sites) for many urban-avoiding reptiles. Rotting wood also improves local soil conditions for fossorial reptiles (MacNally et al., 2001). Fallen wood should therefore be retained in rough and out-of-play areas wherever possible (Table 6.2, 6.3, Fig. 6.3). Many golf courses accumulate wood in large piles. However, these tend to support high densities of urban-adapted reptiles (eg. wall skink: Cryptoblepharus virgatus, eastern water dragon: Physignathus lesueurii and eastern water skink Eulamprus quoyii) and exotic mammals (eg. house mouse: Mus musculus and black rat: Rattus rattus). It is therefore ecologically preferable to retain woody debris in situ.
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Figure 9.3 Retaining dead and hollow-bearing trees will increase habitat availability for arboreal mammals and some birds and reptiles.
Figure 9.4 Retaining native grass in out-of-play areas will provide habitat for mammals, birds and reptiles.
Figure 9.5 Woody debris will support greater reptile diversity if retained in situ.
125 9.2.6 INCREASE AQUATIC VEGETATION COMPLEXITY Golf courses vary dramatically in the extent to which aquatic (floating, emergent and submerged) vegetation is retained in local waterbodies and this in turn, affects the diversity of local amphibian assemblages (Table 6.2, Fig. 6.4). Aquatic vegetation provides critical structural complexity required by larval and post-metamorphic amphibians for food, shelter, feeding, oviposition and calling sites (Beebee, 1981; Verrell, 1987; Raymond and Hardy, 1991; Vos and Chardon, 1998; Hazell et al., 2001). The structural complexity of aquatic vegetation should therefore be enhanced if golf courses are to support regionally threatened amphibians.
9.2.7 DECREASE TURFGRASS COVER One factor universally limiting the abundance and diversity of all vertebrate groups (birds, reptiles, mammals and amphibians) on golf courses was the proportion of turfgrass cover maintained in rough and out-of-play areas (Table 6.2, 6.3). Areas of manicured turfgrass have significantly reduced resource availability and expose local wildlife to increased levels of predation and adverse environmental conditions. Golf courses can increase their conservation value by reducing the extent to which rough and out-of-play areas are devoted to manicured turfgrass. By minimising the area of land that is actively maintained, ongoing maintenance costs (i.e. mowing, irrigation and spraying) can be substantially reduced (AGU, 1998; AGCSA, 1998; Smith, 1998).
9.2.8 NATURALISE WATERBODY DESIGNS Steep-sided, artificial permanent ponds form the majority of waterbodies present on golf courses in SEQ. Amphibian species richness was negatively associated with waterbody bank steepness (Table 6.1, 6.2, Fig. 6.4). Steep banks render waterbodies inaccessible to some amphibian species. The relationship between species richness and bank steepness may however, represent a wider association with co-correlated characteristics likely to be shared by artificial ponds. Ponds with steep banks are also often deep, cement-based and have uniform depth profiles. These factors may reduce the structural diversity of habitats available to larval and post-metamorphic amphibians. Efforts to naturalise waterbody designs by making banks shallower and increasing bottom complexity (i.e. varied depth profiles) are likely to increase the diversity of local amphibian assemblages.
126
Figure 9.6 Natural waterbodies that retain structurally diverse aquatic and riparian vegetation will support greater amphibian diversity than artificial waterbodies with little vegetation.
127 9.2.9 AVOID STOCKING WATERBODIES WITH FISH Other local management actions (not assessed in this study) could have a significant impact on local biodiversity. One prime example is the stocking of waterbodies with exotic fish. Fish predators are a critical determinant of local amphibian assemblage composition. Many amphibian species are excluded from waterbodies where fish are present (Hecnar and M’Closkey, 1997; Hero et al., 1998; 2001). Only those frog species that have evolved antipredator defences (i.e. unpalatability and crypsis) tend to coexist with fish (Wellborn et al., 1996; Hero et al., 1998; 2001). These species are however, generally already common in urban areas (given the high prevalence of fish in urban waterbodies). Fish stocking should therefore be avoided if amphibian conservation is a local management objective.
9.2.10 ALGAL WEED MANAGEMENT: Permanent waterbodies on many golf courses in SEQ experience periodic infestations of the algal weed Hydrilla verticillata. This submerged aquatic macrophyte clogs waterbodies, limiting the growth of other aquatic vegetation, blocking irrigation pumps and restricting water movement (Anderson, 1990; Langeland, 1996). Hydrilla infestations ultimately turn ponds into stagnant, anaerobic pools filled with pungent decomposing vegetation. Various methods are currently used to manage Hydrilla including water- aerating mixers, water-shading dyes, chemical herbicides, benthic barriers and physical removal. Many of these methods are relatively ineffective temporary control measures that may adversely affect local biodiversity by altering local environmental conditions and resource availability and by exposing animals to periodic disturbance. Controlling Hydrilla outbreaks is often difficult and no one strategy is currently universally effective (Langeland, 1996). Hydrilla control should therefore aim to manage the cause of infestations, rather than simply treating the symptoms. Outbreaks of Hydrilla and other aquatic weeds are often exacerbated by elevated nitrogen and phosphorous levels, resulting from fertiliser run-off (AGU, 1998). Effective aquatic weed management may require an integrated management strategy that oxygenates water (using aerating mixers), controls the application of fertilisers (using slow-release fertilisers) and minimises run-off and leaching into waterbodies (by planting vegetated buffer-zones in areas between fertiliser application zones and waterbodies).
128 9.2.11 FACTORS THAT COULD DETER CONSERVATION EFFORTS A number of social and logistical factors could discourage golf clubs from increasing the structural complexity of local habitats. These include health and safety concerns, financial constraints and entrenched expectations regarding appropriate management standards. The most significant constraining factor is the potential to increase fire risk. I) Increased Fire Risk: Efforts to increase the structural complexity of native vegetation on golf courses will inevitably increase fuel loads and thereby elevate the potential fire risk. This threat could however be managed (as it is in many urban bushland areas) through the use of strategic landscape design and controlled prescribed burns. Appropriate fire management is however not simply a health and safety issue, but a critical factor that must be considered in order to maintain local biodiversity. Many native Australian plants and animals have co-evolved with fire and are therefore dependent on periodic fire disturbance and its subsequent regenerative effect, which can promote accelerated nutrient release, seed germination, provide ideal conditions for plant growth and increase the abundance and diversity of resources available to local wildlife (Tran and Wild, 2000; Watson, 2001). Efforts to increase the size and structural complexity of habitats retained on golf courses will therefore have limited conservation value if remnant vegetation is subsequently subjected to inappropriate fire management. Careful consideration is required to find an appropriate balance between hazard reduction and biodiversity management. There is a disparity in the fire frequency required to achieve these goals. In general, low-frequency fire (2-4 years) is required wherever hazard reduction is paramount (i.e. areas close to housing). This will however generally be too frequent to allow sufficient time for many native plants and animals to recover and will therefore tend to reduce biodiversity (Tran and Wild, 2000; Watson, 2001). Recommended fire regimes required to maintain biodiversity in SEQ vary among vegetation types: dry sclerophyll (7-25 years), wet sclerophyll (20-100 years), coastal heathland (7-20 years), Melaleuca (15-30 years) and rainforest (unburnt; Watson, 2001). Wherever possible, these regimes should be used as a goal for land management. As in most urban environments, hazard reduction will be the prime determinant of prescribed burning strategies on golf courses. There will however be greater capacity to implement the prescribed burning regimes recommended to maintain biodiversity if fire hazard reduction principles are incorporated in golf course designs. Careful landscape management can be used to reduce local fire risks, by ensuring large vegetation remnants
129 are retained further from housing, by incorporating strategic fire breaks and utilising permanent ponds and less flammable vegetation types in areas close to housing. This will allow greater freedom to control-burn remaining habitats for biodiversity. Strategies used to promote wildlife conservation will consider not only fire frequency, but the local extent, spatial distribution and season of prescribed burns. A patch-burning or mosaic approach is recommended (i.e. burning different patches in different years) to ensure local wildlife have uninterrupted access to local refuges and to a greater heterogeneity of remnant and regenerating vegetation (Tran and Wild, 2000; Watson, 2001). The season in which burns are conducted can also have a critical influence on wildlife survival and recolonisation success. Varying the season of prescribed burns is generally recommended (Tran and Wild, 2000; Watson, 2001). However, in relatively isolated habitats where local productivity can be critical, it may be advisable to avoid burning in early spring, given their capacity to effect vulnerable juveniles that may have limited dispersal capabilities. Structural features that have critical habitat value (eg. fallen logs and hollows) should be protected from fire by raking surrounding areas prior to prescribed burns (Watson, 2001). Care should also be taken to avoid burning in concentric rings that could trap wildlife, instead ensuring fire-breaks are used to provide wildlife escape routes (Watson, 2001). Prescribed burns should be conducted in conditions of low fire danger, burning from areas of high to low fuel load (Tran and Wild, 2000). Permits are required for any prescribed burn and must be conducted with the approval of local fire wardens. Fire management is a complex issue and the subject of considerable ongoing ecological research. More information regarding appropriate fire management to promote biodiversity can be obtained from the SEQ Fire and Biodiversity Consortium (www.fireandbiodiversity.org.au). A number of other social and logistic factors may constrain efforts to manage golf courses for biodiversity. These can be readily overcome. II) Danger of Tree-fall: Retaining dead and hollow-bearing trees would inevitably increase the potential for injury from falling tree limbs. Dead trees close to playing areas would need to be professionally secured. Warning signs could be erected in areas where dead trees are retained further from playing areas.
130
Figure 9.7 Stocking waterbodies with fish will reduce frog diversity
Figure 9.8 Aquatic weed infestations can be minimised by aerating water and restricting nutrient run-off.
Figure 9.9 Controlled burning is required for fuel reduction and biodiversity management.
131 III) Cost of Planting: Tree planting costs can be minimised by working with government-sponsored community-based tree-planting schemes (eg. Landcare, Land for Wildlife and Greening Australia). By increasing the area of land devoted to complex native vegetation, ongoing mowing, spraying and maintenance costs can also be substantially reduced. IV) Member Backlash: Golf club members may resist changes that affect the perception of their golfing experience. Restoration efforts can therefore only be conducted in a way that does not impinge on the challenge posed by the golf course. There would however be substantial opportunity to increase habitat complexity in out-of-play areas further from playing areas. Education regarding the potential ecological benefits would also help change perceptions and may encourage members to become actively involved in restoration efforts. Wildlife encounters will also greatly enhance the golfing experience. This could have substantial economic value, particularly on tourist-resort golf courses that cater for a potentially lucrative international golf-wildlife market. 9.3 LANDSCAPE SCALE ACTIONS:
IMPLICATIONS FOR GOLF COURSE ARCHITECTS Golf course architects have the greatest capacity to influence the ecological value of suburban golf courses, since the initial golf course design will determine the size and spatial arrangement of vegetation remnants and the structural complexity of local habitats. While ‘natural’ golf course designs have become popular among Australian golf course architects in recent years (Perrett, 1996; Dawson, 2000), this study has shown that there are minimum spatial requirements, below which ‘natural’ designs will have negligible conservation value for threatened wildlife. Aspects of golf course design that are likely to influence local biodiversity include: • area of vegetation (birds, reptiles, amphibians) • patch size (birds, reptiles and mammals) • patch shape (reptiles) • patch connectivity (birds, mammals) • the number of ephemeral waterbodies (amphibians) These factors primarily determine the diversity of ecological niches and resources, the level of shelter provided from predators and environmental threats and the capacity for animals to continue their normal daily and seasonal movements.
132 9.3.1 MAXIMISE THE AREA OF NATIVE VEGETATION RETAINED The most important design characteristic that will affect the refuge-value of individual suburban golf courses is the total area of native vegetation retained in rough and out-of- play areas. The local diversity of threatened birds and amphibians increased significantly with the area of vegetation on golf courses (Fig. 6.1, 6.4) Courses that retained less than 20ha of native vegetation generally had negligible value for threatened birds or amphibians. Golf courses that retain larger areas of native vegetation (40-60ha) inadvertently have a greater abundance and diversity of threatened wildlife. These golf courses will inevitably provide a greater diversity of resources and offer a higher level of protection from predators, noise, disturbance and pollution. This will tend to increase productivity and survivorship in local wildlife populations.
9.3.2 MAXIMISE INDIVIDUAL PATCH SIZES Threatened bird, reptile and mammal diversity increased with vegetation patch size (Chapter 7). Patches smaller than 1-2ha generally had negligible value for threatened species. Small patches often had low structural complexity, and therefore likely to retain few resources and provide little shelter from predators, aggressive competitors, noise and disturbance. In general, patches must be larger than 1-5ha to support high densities of threatened birds and mammals and larger than 5ha to support high densities of threatened reptiles (Fig. 7.4). Golf courses that incorporate larger vegetation patches (>5ha) will inevitably have greater conservation value for threatened birds, reptiles and mammals. There may however be some capacity to increase the ecological value of smaller remnants (1-2ha) by restoring the complexity of understorey and ground-level habitats.
9.3.3 INCREASE PATCH WIDTH Local reptile diversity was associated with patch shape. Many spatially restricted golf courses retain vegetation in narrow linear strips, wedged between fairways. Wildlife in narrow habitats are likely to be exposed to higher levels of predation, competition, noise disturbance and other adverse edge effects (Forman and Godron, 1986; Collinge, 1996). Golf courses with predominantly narrow vegetation patches tend to have low reptile diversity and support few regionally threatened species (Table 6.2, 6.3, Fig. 6.3). Local reptile diversity could be enhanced by retaining wider (>50m) and more rounded vegetation patches. The relative absence of threatened reptiles from narrow patches is also partly attributed to the fact that the understorey of narrow patches is often intensively maintained (since these areas are often close to playing areas). The ecological value of narrow patches (<10m wide) may be enhanced by restoring understorey and ground-level
133 complexity. This would increase on-course connectivity for ground-dwelling animals that cannot readily cross open fairways.
9.3.4 MAXIMISE VEGETATION CONNECTIVITY ON THE COURSE The relative abundance and species richness of threatened mammals was closely associated with the relative connectivity of on-course habitats (Table 6.2, 6.3). This is likely to determine the extent to which shy refuge-dependent mammals including native rodents, the yellow footed antechinus, swamp wallaby and red-necked wallaby can access all parts of the course (including areas that may be connected to external habitats). Birds utilising dense mid-level and understorey vegetation are also likely to benefit from designs that maximise the connectivity of native vegetation.
9.3.5 MAXIMISE THE DIVERSITY OF LOCAL WATERBODY TYPES Amphibian diversity on golf courses is currently limited by a lack of heterogeneity in waterbody types. Artificial permanent ponds form the majority of waterbodies retained on golf courses in SEQ. These provide habitat for a limited number of amphibian species, very few of which are regionally threatened by urbanisation. Ephemeral and semi- permanent waterbodies are generally lost to urbanisation (Rubbo and Kiesecker, 2005). In contrast, permanent waterbodies are often retained for their aesthetic or functional value. Consequently, amphibian species that are dependent on semi-permanent and ephemeral breeding sites are generally declining from urban areas (Rubbo and Kiesecker, 2005). Golf course architects can therefore dramatically increase the local conservation value of golf courses (for amphibians) by ensuring that naturally ephemeral and semi-permanent waterbodies are retained in rough and out-of-play areas (Table 6.2, 6.3 Fig. 6.4). The quality of these waterbodies will also be dependent on the retention of existing native grass and sedges that provide microhabitat complexity necessary for oviposition and larval development.
9.3.6 FACTORS LIMITING THE USE OF ECOLOGICAL DESIGN CRITERIA The extent to which ecological criteria are voluntarily incorporated into golf course designs will vary among golf courses, depending on their type (i.e. whether they have been built as resort, housing or stand-alone golf courses). Few new golf courses are built solely as stand-alone ventures (not associated with any other commercial entity). In SEQ, many new golf courses are now built as part of larger housing developments or
134
Figure 9.10 Maximising the area of vegetation in out-of-play areas will increase bird, reptile, mammal and amphibian diversity.
Figure 9.11 Maximising patch width will increase reptile diversity.
Figure 9.12 Semi-permanent and ephemeral waterbodies support higher
threatened amphibian diversity than permanent ponds.
135 tourist resorts. These courses will have added design constraints that may act as a disincentive to retain native habitat. I) Economic Constraints Housing development golf courses face intense economic pressure to maximise the land area available for housing construction. These golf courses are unlikely to voluntarily retain large areas of native vegetation. Many housing development golf courses that have initially retained areas of native vegetation subsequently clear these areas over time as part of staged housing developments. This significantly reduces their capacity to act as a local wildlife refuge (Chapter 8). Few housing development golf courses in SEQ have substantial conservation value for regionally threatened vertebrates. Stand-alone courses may have a greater capacity to retain native habitat, since it may actually be cheaper to construct courses that conform to the natural contours of the landscape and therefore incur lower engineering and earth-moving costs. II) Limitation imposed by aesthetic expectations of resort golf courses Retaining native Australian vegetation on resort golf courses may require a shift in design philosophy. Resort golf courses around the world employ a common style of landscaping that typically uses tropical vegetation (palms and exotic flowering plants) and manicured turfgrass. These golf courses will have negligible conservation value for threatened Australian vertebrates. Native Australian vegetation has however been incorporated in resort style courses in recent years (eg. the use of Melaleuca and coastal heathland vegetation at Hyatt Regency, Coolum). While the refuge value of this course is uncertain, it demonstrates that native vegetation can be effectively used in resort-style designs. Courses incorporating native vegetation should be used as a benchmark by which to design ecologically valuable resort-style courses and to change expectations regarding the use of native vegetation in golf course designs. III) Hydrological/Engineering constraints Low-lying golf courses close to housing perform a dual recreation/storm-water mitigation role. On these golf courses, large permanent waterbodies act as a source of storm-water mitigation during floods and as a source of irrigation water during drought. While amphibian diversity can be enhanced by increasing the number of ephemeral waterbodies, this should not undermine the functional role performed by existing permanent ponds. Permanent waterbodies with surrounding riparian and aquatic vegetation should therefore be maintained. Efforts should however be made to retain additional semi-permanent and ephemeral waterbodies.
136 9.3.7 USING ‘NATURAL’ COURSE DESIGNS FOR ECONOMIC GAIN There has been considerable movement towards more natural styles of golf course design within the Australian golf industry (Perrett, 1996; Dawson, 2000). There are however, problems associated with this phenomenon. Some golf courses (particularly those associated with housing developments) cosmetically retain native vegetation as an environmental selling point, without having any ecological basis to their design. Some housing development golf courses have minimised the area of native vegetation (to maximise the number of housing lots sold) but strategically retain narrow belts of eucalypt vegetation and charismatic native plants (eg. grass trees and tree ferns) to achieve a ‘natural’ aesthetic without performing any ecological role for threatened native wildlife. This practice has the potential to undermine efforts that have been made in recent decades to improve the golf industries environmental reputation. Greater balance is required between economic and ecological design objectives if golf courses are to realise their conservation potential. Legislation may be required to ensure golf courses protect sufficient areas of native vegetation to support regionally threatened vertebrates. 9.4 REGIONAL SCALE ACTIONS – IMPLICATIONS FOR URBAN PLANNERS While the actions of golf course architects and superintendents will affect the extent to which individual golf courses provide habitat for threatened vertebrate wildlife, it is the actions of local government and urban planners that will ultimately determine the net ecological impact of future golf course developments. By controlling the regional distribution of future golf course development, urban planners and local governments control the type of land that is lost to development and therefore determine the resulting net change in threatened wildlife diversity.
9.4.1 BUSHLAND AREAS SHOULD NOT BE TARGETED FOR GOLF DEVELOPMENT Movement toward more natural styles of golf course design may lead developers to target areas of existing bushland for golf course development. This should be avoided at all costs. The results of this study indicate that threatened wildlife diversity increases with the area of complex native vegetation. Any golf development that involves substantial loss of eucalypt vegetation will therefore lead to a significant localised decline in threatened vertebrate abundance and diversity. Even well-vegetated golf courses are no substitute for continuous stands of eucalypt vegetation. Zoning decisions should therefore discourage golf course development from areas currently occupied by bushland habitat.
137 9.4.2 RESEARCH REQUIRED TO DIRECT ECOLOGICALLY SOUND ZONING DECISIONS More research is required to develop an ecologically sound basis to direct the regional distribution of future golf course development. This thesis has shown that well- vegetated golf courses have a greater capacity to support regionally threatened vertebrates than residential areas. However, while well-vegetated golf courses clearly have refuge value within the context of urban landscapes, it is uncertain whether habitats on golf courses should be actively connected to larger habitat reserves. Information is required on local productivity, survivorship and the nature and intensity of local threats faced by wildlife on golf courses. Population-based research should be conducted on specific threatened species, to assess the potential risks and benefits associated with connecting potentially sub-optimal and optimal habitats in fragmented urban environments. Local threats from predation, noise, disturbance and physical collisions are likely to be relatively low on golf courses. The chemical threats posed by chronic, low-level exposure to herbicides may however be significant and therefore warrant further research. If future research reveals that local productivity and survivorship are relatively high and that local threats are low, golf courses should be actively connected to external habitat networks. Well-vegetated golf courses could then be considered to be a relatively hospitable urban land type that could be linked with low-density urban development zones to buffer existing habitat reserves and create a more cohesive, interconnected network of reserve and off-reserve habitats. However, if golf courses are found to have low productivity and survivorship and high chemical threats from herbicide exposure, habitats on golf courses should be isolated wherever possible from formal reserve networks and drinking-water catchments. Nevertheless, even under these circumstances, well-vegetated golf courses may still have some conservation value, by supporting independent populations of regionally threatened vertebrates.
9.4.3 LEGISLATION TO CONTROL THE ECOLOGICAL VALUE OF GOLF COURSES Urban planners currently have little ability to predict the regional influence that golf courses have on threatened biodiversity. Existing golf courses (that have been built largely without EIA scrutiny or legislative control) display extreme variation in conservation value. Most golf courses retain only small remnants of complex native vegetation and many have vastly simplified understorey complexity. As a result, many golf courses support only common urban-adapted wildlife and have negligible conservation value. While some golf courses support high densities of regionally threatened species, others are less ecologically diverse than residential areas. As such,
138 there is an obvious need for greater regulation governing the retention and use of native vegetation in new golf course developments. Without regulation, the conservation potential identified in this study is unlikely to be realised at an industry-wide level. Golf course designs are increasingly shaped by economic, rather than functional, environmental or even golf-related criteria. This is due to the fact that many new golf courses are now built as part of larger housing developments and therefore face pressure to maximise the area of land available for housing. This will inevitably restrict the capacity to retain native vegetation and will therefore tend to limit the conservation value of suburban golf courses. The golf industries potential conservation value should not be under-estimated. Efforts to increase the size and complexity of wildlife habitats on suburban golf courses will have regional significance in SEQ, particularly given the paucity of lowland habitat, the ongoing development pressure in lowland areas and the current ubiquity of golf courses in the regional lowland landscape. The possible conservation role of suburban golf courses is also of great potential benefit to the golf industry, as this may ensure that golf courses continue to be seen as a legitimate use of land in an urban landscape that is increasingly spatially-restricted. Positive action is however required if the golf industry is to realise its conservation potential. There needs to be greater balance between economic and ecological golf course design objectives. Legislation governing the use of native vegetation in new golf course developments may be required to ensure golf courses retain sufficient areas of complex habitat to provide local refuge for regionally threatened birds, reptiles, mammals and amphibians. If this can be achieved, the golf industry will play a positive ecological role within the urban landscape. By ensuring greater predictability in the ecological value of golf courses (and other open-space areas), urban planners would have a greater capacity to develop ecologically sound urban zoning strategies that optimise the ecological role played by small habitat remnants such as those retained on well-vegetated suburban golf courses.
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171 Appendix 1 – Chapter 2 Classifying reptile, mammal and amphibian species into urban avoiding,
tolerating and exploiting species based on the opinion of six local ecologists. (1 = Less abundant in suburbia than eucalypt forest, 2 = less abundant/comparable, 3 = Comparable abundance, 4 = Comparable/More abundant in suburbia, 5 = More abundant in suburbia than eucalypt forest).
SURVEY RESPONDENT VERTEBRATE GROUP HH LS ND DW SH MH CLASSIFICATION REPTILES Anomalopus verreauxii (Verreaux's skink) 1 2 3 2 Tolerating Calyptotis scutirostrum - Scaly-snouted skink 3 1 1 3 3 1 Tolerating Carlia schmeltzii - Robust Rainbow skink 1 1 1 1 1 Avoiding Carlia vivax - Lively skink 1 1 1 1 1 Avoiding Coerranoscincus reticulatus - Three-toed Snake-tooth skink 1 1 1 1 1 1 Avoiding Cryptoblepharus virgatus - Wall skink 5 5 5 5 5 Exploiting Ctenotus arcanus - No Common Name 1 1 1 1 1 Avoiding Ctenotus robustus - Eastern Striped skink 3 1 3 1 3 1 Tolerating Ctenotus taeniolatus - Copper-tailed skink 1 1 1 1 1 Avoiding Eulamprus martini - Martin's skink 1 1 3 1 1 Avoiding Eulamprus quoyii - Eastern Water skink 4 1 3 3 3 1 Tolerating Eulamprus tenuis - Bar-sided skink 4 1 1 1 1 Avoiding Lampropholis amicula - Secretive skink 1 1 1 1 1 1 Avoiding Lampropholis delicata - Grass skink 5 5 5 5 5 5 Exploiting Ophioscincus truncatus - Short-limbed Snake skink 1 1 1 1 1 1 Avoiding Tiliqua scincoides - Blue-tongue lizard 4 5 3 5 3 1 Tolerating Amphibolurus muricatus - Jacky lizard 1 1 2 1 1 1 Avoiding Diporiphora australis - Eastern Two-lined dragon 1 1 1 1 1 Avoiding Physignathus lesueurii - Eastern Water Dragon 4 1 3 3 3 1 Tolerating Pogona barbata - Bearded Dragon 3 1 1 3 3 1 Tolerating Varanus varius - Lace Monitor 1 1 1 1 1 1 Avoiding Lialis burtonis - Burton's Legless lizard 1 1 1 1 1 1 Avoiding Ramphotyphlops nigrescens - Blind snake 1 1 1 1 1 1 Avoiding Demansia psammophis - Yellow-faced Whip snake 1 1 1 1 1 1 Avoiding Dendrelaphis punctulata - Green Tree snake 1 1 1 2 1 Avoiding Morelia spilota - Carpet Python 3 2 1 1 1 Avoiding Pseudechis porphyriacus - Red-bellied Black snake 1 1 1 1 1 1 Avoiding Rhinoplocephalus nigrescens - Eastern Small-eyed snake 1 1 1 1 1 1 Avoiding Tropidonophis mairii - Keelback snake 1 1 1 1 1 1 Avoiding FROGS Adelotus brevis - Tusked frog 1 1 3 1 2 1 Avoiding Bufo marinus - Cane Toad 5 5 5 5 5 5 Exploiting Crinia parinsignifera - Beeping froglet 3 1 1 3 3 1 Tolerating Crinia signifera - Common Eastern froglet 3 1 1 3 1 1 Avoiding Crinia tinnula - Wallum froglet 1 1 1 1 1 1 Avoiding Limnodynastes dumerilii - Eastern Banjo frog 1 1 3 1 1 1 Avoiding Limnodynastes ornatus - Ornate Burrowing frog 1 3 1 1 1 1 Avoiding Limnodynastes peronii - Striped Marsh frog 5 3 4 5 4 3 Exploiting Limnodynastes salmini - Salmon-striped frog 1 1 1 1 1 1 Avoiding Litoria caerulea - Green Tree frog 1 3 3 3 2 1 Tolerating Litoria dentata - Bleating Tree frog 1 1 1 1 1 1 Avoiding Litoria fallax - Eastern Sedge frog 4 3 3 3 5 3 Tolerating Litoria gracilenta - Dainty Green Tree frog 4 3 1 3 2 1 Tolerating Litoria latopalmata - Broad-palmed frog 1 1 1 1 2 1 Avoiding Litoria nasuta - Rocket frog 1 1 1 1 2 1 Avoiding Litoria peronii - Peron's Tree frog 1 1 1 1 2 1 Avoiding Litoria tyleri - Tyler's Tree frog 1 1 1 1 1 1 Avoiding Mixophyes fasciolatus - Great Barred frog 1 1 1 1 1 1 Avoiding Pseudophryne major - Great Brown Brood frog 1 1 1 1 1 1 Avoiding Pseudophryne raveni - Copper-backed Brood frog 1 1 1 1 1 1 Avoiding Uperoleia fusca - Dusky Toadlet 1 1 1 1 1 1 Avoiding MAMMALS Macropus giganteus - Eastern Grey Kangaroo 1 1 1 1 1 Avoiding Macropus rufogriseus - Red-necked Wallaby 1 1 1 1 1 Avoiding Wallabia bicolor - Swamp Wallaby 1 1 1 1 1 Avoiding Petaurus norfolcencis - Squirrel Glider 1 1 1 1 1 Avoiding Petaurus volans - Greater Glider 1 1 1 1 1 Avoiding Phascolarctos cinereus - Koala 1 1 1 1 1 Avoiding Pseudocheirus peregrinus - Common Ringtail Possum 1 3 5 3 1 Tolerating Trichosurus vulpecula - Common Brushtail Possum 2 3 5 3 1 Tolerating Antechinus flavipes - Yellow-footed Antechinus 1 1 1 1 1 Avoiding Mus musculus - House Mouse 5 5 5 5 5 Exploiting Rattus fuscipes - Bush Rat 1 1 1 1 1 Avoiding Rattus lutreolus - Swamp Rat 1 1 1 1 1 Avoiding Rattus rattus - Black Rat 5 5 5 5 5 Exploiting Rattus tunneyi - Pale Field Rat 1 1 1 1 1 Avoiding
172 Appendix 2 – Chapter 3 Ecological characteristics of individual bird species
Common Name Latin Name Foraging Style Microhabitat Body Size Territoriality Migrant (Cm)
Dollarbird Eurystomus orientalis Hawk Gleaner Canopy 26-30 Non-territorial Gen Leaden Flycatcher Myiagra rubecula Hawk Gleaner Canopy 14-16 Non-territorial FMig Mangrove Gerygone Gerygone laevigaster Hawk Gleaner Canopy 11 Breeding FUnc Rainbow Bee-eater Merops ornatus Hawk Gleaner Canopy 23-28 Breeding Unc Restless Flycatcher Myiagra inquieta Hawk Gleaner Canopy 16-21 Non-territorial FUnc Satin Flycatcher Myiagra cyanoleuca Hawk Gleaner Canopy 15-17 Non-territorial FMig Scarlet Honeyeater Myzomela sanguinolenta Hawk Gleaner Canopy 10-11 Breeding FMig White-throated Gerygone Gerygone olivacea Hawk Gleaner Canopy 10-11.5 Breeding FMig Tree Martin Hirundo nigricans Hawk Gleaner Clear Ground 12.5-14 Non-territorial FUnc Welcome Swallow Hirundo neoxena Hawk Gleaner Clear Ground 15 Non-territorial Clear Willy Wagtail Rhipidura leucophrys Hawk Gleaner Clear Ground 19-22 Breeding Clear Lewin’s Honeyeater Meliphaga lewinii Hawk Gleaner Closed Mid-storey 19-21.5 Permanent FRes Rufous Whistler Pachycephala rufiventris Hawk Gleaner Closed Mid-storey 17-18 Non-territorial FMig Spangled Drongo Dicrurus bracteatus Hawk Gleaner Closed Mid-storey 29-32 Non-territorial FUnc Fan-tailed Cuckoo Cuculus flabelliformis Hawk Gleaner Undergrowth 25-27 Non-territorial FUnc Grey Fantail Rhipidura fuliginosa Hawk Gleaner Undergrowth 14-17 Non-territorial FMig Rufous Fantail Rhipidura rufifrons Hawk Gleaner Undergrowth 15-16.5 Non-territorial FMig Bellbird Manorina melanophrys High Gleaner Canopy 19 Permanent FRes Black-faced Cuckoo-shrike Caracina novaehollandiae High Gleaner Canopy 30-36 Breeding Gen Brown Thornbill Acanthiza pusilla High Gleaner Canopy 10 Non-territorial FRes Brown Treecreeper Climacteris picumnus High Gleaner Canopy 16-18 Permanent FRes Brush Cuckoo Cuculus variolosus High Gleaner Canopy 22-26 Non-territorial FUnc Channel-billed Cuckoo Scythrops novaehollandiae High Gleaner Canopy 58-66 Non-territorial FMig Cicadabird Coracina tenuirostris High Gleaner Canopy 24-26 Breeding FUnc Common Koel Eudynamis scolopacea High Gleaner Canopy 40-46 Non-territorial Gen Figbird Sphecotheres viridis High Gleaner Canopy 27-30 Non-territorial Unc Horsfield’s Bronze Cuckoo Chrysoccyx basalis High Gleaner Canopy 15-17 Non-territorial Unc Large-billed Scrubwren Sericornis magnirostris High Gleaner Canopy 12-13 Non-territorial FRes Striated Pardalote Pardalotus striatus High Gleaner Canopy 9.5-12 Non-territorial FMig Weebill Smicrornis brevirostris High Gleaner Canopy 8-11 Breeding FRes White-throated Honeyeater Melithreptus albogularis High Gleaner Canopy 13-15 Non-territorial FUnc White-throated Treecreeper Cormobates leucophaea High Gleaner Canopy 16-17.5 Non-territorial FRes Yellow-faced Honeyeater Lichenostomus chrysops High Gleaner Canopy 16-18 Breeding FMig Brown Honeyeater Lichmera indistincta High Gleaner Closed Mid-storey 11-15 Breeding Gen Olive-backed Oriole Oriolus sagittatus High Gleaner Closed Mid-storey 26-28 Non-territorial Gen Silvereye Zosterops lateralis High Gleaner Closed Mid-storey 10-12.5 Non-territorial Gen Mistletoe-bird Dicaecum hirundinaceum High Gleaner Closed Mid-storey 10-11 Non-territorial FUnc Little Friarbird Philemon citreogularis High Gleaner Open Mid-storey 25-29 Permanent Gen Little Wattlebird Anthochaera lunulata High Gleaner Open Mid-storey 26-31 Non-territorial Unc Noisy Friarbird Philemon corniculatus High Gleaner Open Mid-storey 31-35 Breeding Gen Rainbow Lorikeet Trichoglossus haematodus High Gleaner Open Mid-storey 25-32 Non-territorial Gen Scaly-breasted lorikeet Trichoglossus chlorolepidotus High Gleaner Open Mid-storey 22-24 Non-territorial Clear
FMig = Forest Migrants, FRes = Forest Residents, FUnc = Forest Uncertain, Unc = Uncertain, Gen = Generalist, Clear = Cleared Land (Categories from Bentley and Catterall, 1997).
173 Appendix 3 – Chapter 3 Ecological characteristics of individual bird species
Latin Name Common Name Foraging Style Microhabitat Body Size Territoriality Migrant (Cm)
Crested Pigeon Ocyphaps lophotes Low Granivore Clear Ground 0-34 Non-territorial Gen Feral Pigeon Columba livia Low Granivore Clear Ground 33-36 Non-territorial Clear Galah Cacatua roseicapilla Low Granivore Clear Ground 34-38 Non-territorial Clear House Sparrow Passer domesticus Low Granivore Clear Ground 14-16 Non-territorial Clear Little Corella Cacatua sanguinea Low Granivore Clear Ground 35-39 Non-territorial Clear Spotted Turtledove Streptopelia chinensis Low Granivore Clear Ground 28-32 Non-territorial Clear Sulphur-crested Cockatoo Cacatua galerita Low Granivore Clear Ground 44-51 Non-territorial Gen Bar-shouldered Dove Geopelia humeralis Low Granivore Closed Mid-storey 26-30 Non-territorial Unc Peaceful Dove Geopelia striata Low Granivore Closed Mid-storey 19-21 Non-territorial Unc Red-browed Firetail Neochmia temporalis Low Granivore Closed Mid-storey 11-11.5 Non-territorial FRes Double-barred Finch Poephila bichenovii Low Granivore Undergrowth 10-11 Non-territorial Gen Noisy Miner Manorina melanocephala Low Omnivore Canopy 24-27 Permanent Clear Pied Currawong Strepera graculina Low Omnivore Canopy 42-50 Permanent Unc Common Myna Acridotheres tristis Low Omnivore Clear Ground 23-25 Permanent Clear Common Starling Sturnus vulgaris Low Omnivore Clear Ground 21 Non-territorial Clear Magpie-lark Grallina cyanoleuca Low Omnivore Clear Ground 26-30 Permanent Clear Pheasant Coucal Centropus phasianinus Low Omnivore Closed Mid-storey 50-70 Non-territorial Gen Grey-crowned Babbler Pomatostomus temporalis Low Omnivore Closed Mid-storey 25 Non-territorial Unc Blue-faced Honeyeater Entomyzon cyanotis Low Omnivore Open Mid-storey 30-32 Permanent Gen Pale-headed Rosella Platycercus adscitus Low Omnivore Open Mid-storey 28-32 Non-territorial Gen Australian Brush-turkey Alectura lathami Low Omnivore Undergrowth 60-70 Breeding Unc Australian King Parrot Alisterus scapularis Low Omnivore Undergrowth 41-44 Non-territorial Unc Eastern Whipbird Psophodes olivaceus Low Omnivore Undergrowth 25-31 Permanent FRes Golden-headed Cisticola Cisticola exilis Low Omnivore Undergrowth 9-11.5 Breeding Gen Painted Button Quail Turnix varia Low Omnivore Undergrowth 17-19 Breeding FRes Red-backed Wren Malurus melanocephalus Low Omnivore Undergrowth 10-13 Breeding FRes Spotted Quail Thrush Cinclosoma punctatum Low Omnivore Undergrowth 25-28 Non-territorial Unc Superb Blue Wren Malurus cyaneus Low Omnivore Undergrowth 13-14 Permanent Gen Variegated wren Malurus assimilis Low Omnivore Undergrowth 11-14.5 Breeding FRes White-browed Scrubwren Sericornis frontalis Low Omnivore Undergrowth 11-13 Breeding FRes Chestnut-breasted Mannikin Lonchura castaneothorax Low Omnivore Undergrowth 10 Non-territorial Gen Forest Kingfisher Todiramphus macleayii Pounce Gleaner Canopy 18-23 Breeding FUnc Sacred Kingfisher Todiramphus sancta Pounce Gleaner Canopy 20-23 Breeding FMig Australian Magpie Gymnorhina tibicen Pounce Gleaner Clear Ground 38-44 Permanent Clear Torresian Crow Corvus orru Pounce Gleaner Clear Ground 48-53 Permanent Gen Eastern Yellow Robin Eopsaltria australis Pounce Gleaner Closed Mid-storey 15-16 Non-territorial FRes Grey Shrike-thrush Colluricincla harmonica Pounce Gleaner Closed Mid-storey 22-26 Breeding FRes Laughing Kookaburra Dacelo noveguineae Pounce Gleaner Closed Mid-storey 41-47 Permanent Gen Grey Butcherbird Cracticus torquatus Pounce Gleaner Open Mid-storey 24-30 Permanent Gen Mangrove Kingfisher Todiramphus chloris Pounce Gleaner Open Mid-storey 24-29 Breeding FUnc Pied Butcherbird Cracticus nigrogularis Pounce Gleaner Open Mid-storey 32-36 Permanent Gen Buff-rumped Thornbill Acanthiza reguloides Pounce Gleaner Undergrowth 11 Non-territorial FUnc Golden Whistler Pachycephala pectoralis Pounce Gleaner Undergrowth 17-19 Non-territorial FMig
FMig = Forest Migrants, FRes = Forest Residents, FUnc = Forest Uncertain, Unc = Uncertain, Gen = Generalist, Clear = Cleared Land (Categories from Bentley and Catterall, 1997).
174
Appendix 4 – Chapter 3 Ecological characteristics of individual reptile species
Table 3.2 Ecological characteristics of individual reptile species
Species Common Name Body Microhabitat Taxonomy Size (cm)
Diporiphora australis Jacky Lizard 50 Arboreal Agamid Cryptoblepharus virgatus Wall Skink 40 Arboreal Skink Dendrelaphis punctulata Common Tree Snake 1200 Arboreal Snake Ctenotus robustus Robust Skink 110 Complex Ground Cover Skink Lampropholis amicula Secretive Skink 30 Complex Ground Cover Skink Carlia vivax Lively Skink 40 Complex Ground Cover Skink Carlia schmeltzii Shmeltz’s Skink 50 Complex Ground Cover Skink Eulamprus martini Scincidae 70 Complex Ground Cover Skink Ctenotus taeniolatus Scincidae 80 Complex Ground Cover Skink Ctenotus arcanus Scincidae 90 Complex Ground Cover Skink Coeranoscincus reticulatus. Reticulated Skink 180 Complex Ground Cover Skink Demansia psammophis Eastern Small-Eyed Snake 700 Complex Ground Cover Snake Pseudechis porphyriacus Red-Bellied Black Snake 1500 Complex Ground Cover Snake Varanus varius Lace Monitor 2000 Complex Ground Cover Varanid Lialis burtonis Burton’s Legless Lizard 250 Fossorial Pygopods Calyptotis scutirostrum Skink 45 Fossorial Skink Anomalopus verreauxii Verreaux’s Skink 150 Fossorial Skink Rhinoplocephalus nigrescens Elapidae 1200 Fossorial Snake Ramphotyphlops nigrescens Typhlopidae 380 Fossorial Typhlopids Physignathus lesueuri Eastern Water Dragon 2000 Semi-aquatic Agamid Eulamprus quoyii Eastern Water Skink 100 Semi-aquatic Skink Tropidonophis mairii Keelback Snake 500 Semi-aquatic Snake Pogona barbata Bearded Dragon 250 Simple Ground Cover Agamid Lampropholis delicata Grass Skink 40 Simple Ground Cover Skink Morelia spilota Carpet Python 3000 Simple Ground Cover Snake
All data from Cogger, H.G. (1996) Reptiles and Amphibians of Australia. Reed Books Australia.
175 Appendix 5 – Chapter 3 Ecological characteristics of individual amphibian species Table 3.3 Ecological characteristics of amphibian species
Common Name Species Body Size Breeding Larval Clutch size Microhabitat (Cm) Site Period (# Eggs) (Days) Tusked Frog Adelotus brevis 45 Permanent 50A 638B Aquatic Cane Toad Bufo marinus 150 Generalist 30M 25000M Terrestrial Beeping Froglet Crinia parinsignifera 20 Ephemeral 164A 130C Sedge Common Eastern Froglet Crinia signifera 30 Generalist 49D 216C Generalist Wallum Froglet Crinia tinnula 30 Ephemeral <180A Unknown Sedge Eastern Banjo Frog Limnodynastes dumerilii 70 Permanent 180A 2348C Aquatic Ornate Burrowing Frog Limnodynastes ornatus 45 Permanent 90E 759F Terrestrial Striped Marsh Frog Limnodynastes peronii 65 Permanent 270A 1319G Aquatic Salmon-striped Frog Limnodynastes salmini 75 Permanent 43H 1630H Aquatic Green Tree Frog Litoria peronii 50 Permanent 105A 1777C Arboreal Bleating Tree Frog Litoria caerulea 100 Permanent 40A 2753F Arboreal Eastern Sedge Frog Litoria dentata 45 Permanent 51A 1070G Arboreal Dainty Green Tree Frog Litoria fallax 25 Permanent 132K 941F Floating veg. Broad Palmed frog Litoria gracilenta 45 Permanent 60A Unknown Arboreal Rocket Frog Litoria latopalmata 40 Permanent Unknown 336A Terrestrial Peron's Tree Frog Litoria nasuta 50 Permanent 53L 28A Terrestrial Tyler's Tree Frog Litoria tyleri 50 Permanent 80A 128G Arboreal Great Barred Frog Mixophyes fasciolatus 80 Permanent >365A 977I Terrestrial Great Brown Brood Frog Pseudophryne major 40 Ephemeral 210A Unknown Sedge Copper-backed Brood Frog Pseudophryne raveni 40 Ephemeral 47*J 74*J Sedge Dusky Toadlet Uperoleia fusca 30 Ephemeral 90A Unknown Sedge
A. Anstis, M. (2002). Tadpoles of south-eastern Australia. New Holland Publishers, Australia B. Daly, G. (1995). Observations on the Tusked Frog, Adelotus brevis (Anura: Myobatrachidae). Herpetofauna 25 (2): 32-35. C. Humphries, R.B., (1979) Dynamics of a breeding frog community. PhD thesis, Australian National University, Canberra. D. Moore, J.A. (1961) The frogs of eastern New South Wales. Bulletin of the American Museum of Natural History, 121: 149-386. E. Tyler, M.J., Crook, G.A., and Davies, M. (1983) Reproductive Biology of the frogs of the Magela Creek system, Northern Territory. Records of the South Australian Museum, 18 (18): 415-440. F. Dziminski, M. (2000) Intraclutch variation in egg provisioning in typical Australian frogs: its consequences for larval ecology. Honours thesis, James Cook University, Townsville. G. Daly, G. (1997) Unpublished data. H. Davies, M. and Watson, G.F. (1994) Morphology and reproductive biology of Limnodynastes salmini, L. convexiusculus and Megistolotis lignarius (Anura: Leptodactylidae: Limnodactylinae). Transactions of the Royal Society of South Australia, 118 (3): 149-169. I. Morrison, C. and Hero J-M. (200 ) Geographic variation…. J. Thumm, K. (1999) Unpublished data. K. Watson, G.F. and Martin, A.A. (1979) Early development of the Australian green hylid frogs Litoria chloris, L. fallax and L. gracilenta. The Australian Zoologist, 20(2): 259-268. L. Jenkinson, D. (1999) Pers. comm. M. Tyler, M.J. (1994) Australian Frogs, A natural history. Reed New Holland, Sydney.
176 Appendix 6 – Chapter 3 Table 3.4 Ecological characteristics of individual mammal species Ecological characteristics of individual mammal species
Species Common Name Size (g) Habit Diet Home Max. Specialisation Range (Ha) Young/Year
Mus Musculus House Mouse 10 Ground Generalist 0.01J, K 60L Rattus rattus Black Rat 280 Ground Generalist 0.01 60 Pseudocheirus peregrinus Common Ringtail 1000 Arboreal Generalist 0.33 – 2.6E 6B Trichosurus vulpecula Common Brushtail 2000 Arboreal Generalist 2.4 – 5.4F 2.5F Phascolarctos cinereus Koala 12000 Arboreal Specialist 5 - 12.5C 1B Petaurus norfolcensis Squirrel Glider 230 Arboreal Specialist 3 – 5D 1.5B Petaurus volans Greater Glider 1200 Arboreal Specialist 0.7 – 3D 1B Antechinus flavipes Yellow-footed Antechinus 56 Ground Generalist 0.3A 8B Macropus giganteus Eastern Grey Kangaroo 50000 Ground Generalist 12.7 – 36.8G 1B Macropus rufogriseus Red-necked Wallaby 16000 Ground Specialist 15.2 – 31.6H 1.4B Wallabia bicolor Swamp Wallaby 14000 Ground Specialist 4.5 – 7.2I 1 Rattus fuscipes Bush Rat 140 Ground Generalist 0.1-0.4L 3L Rattus lutreolus Swamp Rat 100 Ground Specialist 3L 3L Rattus tunneyi Pale Field Rat 12 Ground Specialist 1.5L 11L
A. Smith, G.C. (1984). The Biology of the Yellow-footed Antechinus, Antechinus flavipes (Marsupalia: Dasyuridae), in a Swamp Forest on Kinaba Island, Cooloola, Queensland. Australian Wildlife Research, 11: 465-480. B. Lee, A.K. and Ward, S.J. (1989) Life histories of macropodoid marsupials. Pages 105-115, In Grigg, G., Jarman, P. and Hume, I. (eds). Kangaroos, wallabies and rat-kangaroos. Surrey Beatty and Sons, Sydney, Australia. C. White, N.A. (1999) Ecology of the koala (Phascolarctos cinereus) in rural south-east Queensland, Australia. Wildlife Research, 26: 731-744 D. Lindenmayer, D.B. (2002) Gliders of Australia, A natural history. University of New South Wales Press. E. Kerle, (2001) Possums The Brushtails, Ringtails and Greater Glider. University of New South Wales Press Ltd. F. Green, W.Q. (1984) A review of ecological studies relevant to the management of the Common Brushtail Possum. In Smith, A. and Hume, I. (Eds) Possums and Gliders,Surrey Beatty and Sons, Sydney, Australia. G. Jaremovic and Croft (1991) Social organization of the eastern grey kangaroo (Macropodidae, Marsupalia) in southeastern New South Wales: 1. Groups and group home ranges. Mammalia 55: 169-185 H. Johnson, C.N. (1987) Macropod Studies at Wallaby Creek. IV. Home Range and Movements of the Red-Necked Wallaby. Australian Wildlife Research, 14: 125-132. I. Edwards, G.P. and Ealey, E.H.M. (1971) Aspects of the Ecology of the Swamp Wallaby Wallabia bicolour (Marsupalia: Macropodidae). Australian Mammalogy, 1: 307-317. J. Mikesic, D.G. and Drickamer, L.C. (1992) Factors affecting home range size in house mice (Mus musculus domesticus) living in outdoor enclosures. American Midland Naturalist, 127: 31-40. K. Zielinski, W.J., vom Saal, F.S. and Vandenbergh, J.G. (1992) The effect of intrauterine position on the survival, reproduction and home range size of female house mice (Mus musculus). Behavioural Ecology and Sociobiology. 30: 185-191. L. Watts, C.H.S. and Aslin, H.J. (1981) Rodents of Australia. Angus and Robertson Publishers, Australia.
177 Appendix 7 – Chapter 7 TableOccupancy 7.1a Occupancy rates of bird rates species of bird in differentspecies in patch different size categoriespatch size categories.
Patch size categories Species 0-0.1ha 0.1-0.5ha 0.5-1ha 1-5ha >5ha Forest n = 30 n = 95 n = 41 n = 49 n = 5 n = 10
Noisy Miner 77 91 93 71 80 25 Rainbow Lorikeet 55 72 76 80 100 73 Crested Pigeon 52 43 44 39 60 3 Pied Butcherbird 48 49 46 49 80 37 Magpie-lark 45 35 37 47 60 17 Black-backed Magpie 35 38 54 51 100 48 Willy Wagtail 26 19 29 51 40 2 Torresian Crow 23 36 51 76 80 59 Figbird 23 20 22 45 40 6 Common Myna 23 18 15 22 0 2 Scaly-breasted Lorikeet 16 15 32 29 40 8 Galah 16 9 15 20 80 5 Black-faced Cuckoo-shrike 13 3 20 24 20 39 Spotted Turtle-Dove 10 22 22 43 0 15 Little Corella 10 13 5 6 0 1 Pale-headed Rosella 10 9 17 24 40 33 Sulphur Crested Cockatoo 10 6 10 16 0 6 Grey Butcherbird 6 31 39 55 60 36 Brown Honeyeater 6 11 5 29 40 15 Whistling Kite 6 3 2 0 20 5 Olive-backed Oriole 6 2 12 39 80 40 Pied Currawong 3 12 22 12 20 31 Welcome Swallow 3 12 7 6 0 0 Striated Pardalote 3 7 20 45 100 71 Dollarbird 3 4 5 33 80 14 Bar-shouldered Dove 3 4 5 16 80 15 Sacred Kingfisher 3 2 10 41 60 29 Little Friarbird 3 1 2 4 20 0 Tawny Frogmouth 3 1 0 4 0 1 Blue-faced Honeyeater 3 1 2 0 0 0 Laughing Kookaburra 0 7 10 39 60 29 Common Koel 0 7 2 20 20 3 Superb Blue Wren 0 5 7 29 40 11 Noisy Friarbird 0 5 12 20 60 48 Silvereye 0 3 2 20 60 23 Grey Shrike-thrush 0 2 7 37 100 67 White-throated Honeyeater 0 2 0 22 100 50 Rufous Whistler 0 2 0 20 40 29 Variegated Wren 0 2 2 18 40 9 Channel-billed Cuckoo 0 2 7 6 60 5 Double-barred Finch 0 2 0 6 0 9 Collared Kingfisher 0 2 5 4 0 0 Large-billed Scrubwren 0 2 0 2 20 3 Chestnut-breasted Mannikin 0 2 0 2 0 0 Feral Pigeon 0 1 0 2 0 0 House Sparrow 0 0 0 2 0 0 Long-billed Corella 0 0 2 0 0 0
178 Appendix 7 – Chapter 7 Occupancy rates of bird species in different patch size categories
Patch size categories Species 0-0.1ha 0.1-0.5ha 0.5-1ha 1-5ha >5ha Forest n = 30 n = 95 n = 41 n = 49 n = 5 n = 10
Eastern Whipbird 0 1 2 29 20 25 White-throated Gerygone 0 0 5 20 60 40 Spangled Drongo 0 0 0 20 60 21 Brown Thornbill 0 0 0 18 60 17 Lewin's Honeyeater 0 1 0 18 20 11 Pheasant Coucal 0 0 5 18 20 3 White-throated Treecreeper 0 0 0 14 20 45 Grey Fantail 0 0 0 14 20 36 Scarlet Honeyeater 0 1 0 12 80 60 Golden Whistler 0 0 0 10 100 28 Cicadabird 0 1 0 10 60 20 Rainbow Bee-eater 0 1 2 8 80 29 Fantailed Cuckoo 0 1 0 0 40 10 Forest Kingfisher 0 0 0 8 60 6 Satin Flycatcher 0 0 0 6 20 2 Rufous Fantail 0 0 0 6 20 10 Peaceful Dove 0 1 0 4 20 12 Yellow-faced Honeyeater 0 0 0 12 0 24 Australian Brush-turkey 0 1 0 12 0 0 Red-backed Wren 0 1 0 8 0 12 Horsfield's Bronze Cuckoo 0 0 2 8 0 4 Little Wattlebird 0 0 0 8 0 3 White-browed Scrubwren 0 0 2 4 0 7 Restless Flycatcher 0 0 2 4 0 5 Eastern Yellow Robin 0 0 0 4 0 5 Mangrove Gerygone 0 1 5 4 0 3 Spotted Pardalote 0 0 0 4 0 0 Grey Crowned Babbler 0 0 0 2 40 0 Eastern Rosella 0 1 0 2 20 0 Australian King Parrot 0 1 0 0 20 4 Red-browed Firetail 0 2 0 0 20 2 Brown Cuckoo-dove 0 0 0 0 20 2 Leaden Flycatcher 0 0 0 2 0 6 Brown Treecreeper 0 0 0 2 0 3 Weebill 0 0 0 2 0 3 Mistletoebird 0 0 2 2 0 1 Collared Sparrowhawk 0 0 0 2 0 1 Brush Cuckoo 0 0 0 2 0 1 Azure Kingfisher 0 1 0 2 0 0 Pacific Baza 0 0 0 2 0 0 Southern Boobook 0 0 0 2 0 0 Varied Triller 0 0 0 2 0 0 Painted Button Quail 0 0 0 0 4 0 Buff-rumped thornbill 0 0 0 0 0 10 Golden-headed Cisticola 0 1 0 0 0 1 Glossy Black Cockatoo 0 0 0 0 0 1 Powerful Owl 0 0 0 0 0 1 Brown Falcon 0 1 0 0 0 0 Tree Martin 0 1 0 0 0 0
179 TableAppendix 7.2 Occupancy 8 Occupancy rates rates of reptile of reptiles in differentspecies in patch different size patchcategories sizes
Species 0-0.1ha 0.1-0.5ha 0.5-1ha 1-5ha >5ha Forest n = 30 n = 95 n = 41 n = 49 n = 5 n = 10 REPTILES Tiliqua scincoides 0 0 0 2 0 2 Physignathus lesueurii 32 22 15 39 0 7 Lampropholis delicata 29 46 76 80 100 85 Cryptoblepharus virgatus 26 32 44 39 100 66 Pogona barbata 13 9 17 10 40 12 Eulamprus quoyii 3 1 0 6 0 0 Dendrelaphis punctulata 6 3 2 6 0 3 Calyptotis scutirostrum 3 4 24 24 0 12 Anomalopus verreauxii 3 0 0 0 0 4 Lampropholis amicula 0 2 10 4 40 18 Carlia vivax 0 2 5 2 20 49 Ctenotus taeniolatus 0 1 5 4 20 9 Varanus varius 0 1 2 4 20 24 Diporiphora australis 0 1 0 0 0 1 Pseudonaja textilis 0 1 0 0 0 0 Demansia psammophis 0 0 0 0 0 1 Ctenotus robustus 0 0 0 0 20 4 Eulamprus martini 0 0 5 2 20 11 Lialis burtonis 0 0 5 2 0 5 Carlia schmeltzii 0 0 0 2 0 2 Ctenotus arcanus 0 0 0 0 0 2 Tropidonophis mairii 0 0 0 2 0 0 Morelia spilota 0 0 0 2 20 0
Ophioscincus truncatus 0 0 0 2 0 0 Ramphotyphlops nigrescens 0 0 0 0 0 6 Eulamprus tenuis 0 0 0 0 0 3 Rhinoplocephalus nigrescens 0 0 0 0 0 2 Amphibolurus muricatus 0 0 0 0 0 2 Coerranoscincus reticulatus 0 0 0 0 0 1 Pseudechis porphyriacus 0 0 0 0 0 1
Appendix 9 Occupancy rates of mammal species in different patch sizes
MAMMALS Black Rat 3 5 10 16 0 0 House Mouse 0 8 2 2 0 1 Common Ringtail Possum 10 5 10 0 0 3 Common Brushtail Possum 7 7 7 35 20 22 Grey Kangaroo 7 1 0 12 60 1 Swamp Rat 3 1 0 2 0 1 Koala 0 3 5 2 0 8 Bush Rat 0 1 0 4 60 4 Squirrel Glider 0 1 0 2 0 4 Fox 0 0 2 0 0 2 Red-necked Wallaby 0 0 2 6 60 7
Yellow-footed Antechinus 0 0 2 2 0 7 Pale Field rat 0 0 0 2 0 1 Swamp Wallaby 0 0 0 2 20 0 Greater Glider 0 0 0 0 0 3 Wild Dog 0 0 0 0 0 1
180