Barriers to fish migration in drying climates: contributions from south-western

Mark Gregory Allen

This thesis is presented for the Doctor of Philosophy

2016

Murdoch University, ,

i

DECLARATION

I declare that this thesis is my own account of my research (except as indicated below) and contains as its main content work that has not previously been submitted for a degree at any tertiary education institution

Mark Gregory Allen

THESIS STRUCTURE

This thesis has been compiled as discrete papers for publication, most of which were co- authored by my principal supervisors and other collaborators as indicated below:

Chapters 1 and 6: Mark Allen (100%) Chapter 2: Mark Allen (80%), Stephen Beatty (15%), Alan Lymbery (5%) Chapter 3: Mark Allen (95%), Stephen Beatty (5%) Chapter 4: Mark Allen (90%), Stephen Beatty (10%) Chapter 5: Mark Allen (85%), Stephen Beatty (15%)

CO-AUTHOR’S DECLARATION

As co-authors of the aforementioned works, we hereby give our consent to their inclusion in the thesis.

Mark G. Allen Stephen J. Beatty Alan J. Lymbery

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Acknowledgements

Many people have inspired and assisted me over the course of completing this thesis, but first and foremost, I wish to express my sincerest gratitude to my supervisors Dr David

Morgan and Dr Stephen Beatty. I will be forever in their debt for bringing me into the fold of the Freshwater Fish Group & Fish Health Unit (FFGFHU) and making me feel like an integral and valued team member. Their work has been, and continues to be, an inspiration, and without their guidance and encouragement this thesis never would have eventuated. Above all else, I wish to thank them for their friendship and for all the good times shared over the past 15 years. I already miss the field trips, the bush, the banter, the

Bollies, the Muddies, the Trutts, and the LPPs. I look forward to our next collaboration with great anticipation….and yes it’s definitely my buy. Cheers fellas!

To my loving family, I offer my sincerest thanks for their unwavering support of my work and for the patience and understanding that they have shown despite me having spent far too much time away from them chasing fish. With the thesis now finished, it’s finally time for us to press on with all the plans that have been on hold for what seems like an eternity. I look forward to all the exciting times that lie ahead.

Special thanks are also reserved for my parents, whom I cannot thank enough for their constant love and support and for raising me to have an appreciation and fascination of the natural world. I cannot think of a better role model in life than my mother Connie, or a better career mentor than my father Gerry. I’m not the first person to have been inspired by my father to pursue a career in ichthyology, nor will I be the last. Following in his footsteps was always ‘on the cards’ for me and it has been a fantastic and rewarding journey thus far. Long may it continue for the both of us! iii

My good friend Rob Jurjevich is thanked for his technical assistance with the project, and more importantly, for his positivity, encouragement, generosity, humour and friendship, which has been a tonic and a joy over the years.

I also wish to make special mention of the assistance provided by James Keleher, whose tireless work ethic and dedication to mastering whatever task he is presented with is unsurpassed. I wish him the very best in his future endeavours. The following people are also thanked for their assistance in the completion of this work: Alan Lymbery,

Gillian White, John Snowball, Mark Adams, Gordon Thomson, Paul Close, Tom Ryan,

Ken & Jan Beatty, Trine Beatty, Iain Rankin, Belinda Robson, Brent Nottage, “Debo”

(the chopper pilot), Harry Dunkley, Ron McLean, Amy Mutton, Kath Lynch, Tim Storer,

Tim Marsden, Dave Bloomfield, Andrew Maughan, Steve Childs, George Doust, Jeff

Whitty, Lynn Heppell, Mark Waud, Chris Sharpe, Rob Freeman, Michelle Miller,

Richard Stewart, Brian Stockwell, Nick Bond, John Koehn, Paul Duncanson, Craig

Lawrence, Debra Thomas, Deon Utber, Adrian Stratico, Judy Maughan, Joslyn

Berkelaar, Pip Marshall, Emily Hugues-dit-Ciles, Ash Ramsey, Michael Klunzinger, and the late Jon Murphy. I also wish to thank the journal editors, anonymous reviewers and examiners who provided helpful feedback on this work.

The following departments, agencies and organisations are thanked for their role in facilitating the completion of various aspects of the project: Centre for Fish &

Fisheries Research (Murdoch University), Department of Water (Government of Western

Australia), Department of Fisheries (Government of Western Australia), Department of

Parks and Wildlife (Government of Western Australia), Water Corporation of Western

Australia, Centre of Excellence in Natural Resource Management (University of Western

Australia), Geocatch, South Coast Natural Resource Management, South West

Catchments Council, Oyster Harbour Catchment Group, Wilson Inlet Catchment iv

Committee, Blackwood Basin Group, Cape to Cape Catchments Group, and the Warren

Catchments Council.

This study complied with scientific permits granted by the Department of

Fisheries (Government of Western Australia), Department of Parks and Wildlife

(Government of Western Australia), and the Murdoch University Ethics

Committee (Permit No. RW2579/13).

I would like to acknowledge and thank the following funding bodies that made this research possible: the National Climate Change Adaptation Research Facility

(Commonwealth Government of Australia), the State NRM Office of Western Australia, the Water Corporation of Western Australia, and the Australian Government’s National

Landcare Program.

Finally, I acknowledge the Nyoongar people who are the Traditional Custodians of the Land on which this research took place.

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Abstract

Threats to the depauperate and highly endemic freshwater fish fauna of south-western

Australia have escalated in recent decades, leading to the listing of five of the eleven

(~45%) teleost species as threatened or endangered under various Acts of State and

Federal legislation. The most profound impacts on these fishes have resulted from salinisation of waterways, riparian habitat degradation, proliferation of instream barriers, and the introduction of exotic fish species. In addition, annual rainfall has declined considerably in south-western Australia since the 1970s, resulting in a dramatic reduction in annual surface flow and groundwater levels. This trend is projected to continue and is likely to exacerbate other stressors and directly impact upon the region’s unique freshwater fishes. Using south-western Australia as a model region, this study aimed to investigate how a drying climate may influence the impacts of instream barriers on freshwater fish migration, and importantly, how it may influence management actions designed to mitigate these impacts. In achieving this aim, the study also addressed key knowledge gaps on distributional range, seasonal movements and life history traits of two of the region’s most endangered fishes: the Trout Minnow ( truttaceus) and the recently described Little Pygmy Perch (Nannoperca pygmaea).

A global review of instream barrier impacts revealed a wide range of effects on aquatic ecosystems such as loss of habitat, alteration of natural flow regimes, and disruption of longitudinal river connectivity. Attempts to mitigate their effects on migratory fishes have most commonly involved fishway construction; however, complete physical removal of redundant barriers is increasing globally (particularly in the United

States), mainly due to safety concerns and prohibitive costs of maintaining aging structures. The review revealed a number of processes used in different jurisdictions for vi assessing and prioritising barriers for remediation or removal; however, the influence of climate change was not specifically considered in any process.

Physico-chemical shifts in river systems, reductions in surface discharge and alterations to seasonal flow regimes driven by climate change are likely to exacerbate the impacts of instream barriers on freshwater fishes; however, the manner in which such changes will influence the mitigation of impacts of instream barriers has received limited attention. In south-western Australia, climate change will exacerbate instream barrier impacts by reducing habitat connectivity, compromising the efficacy of fishways, facilitating the introduction of climatically mismatched species, and adding pressure on environmental water allocations. Paradoxically, climate change may also enhance the conservation value of some barriers and their associated impoundments as ecological refuges. The trade-offs between the positive and negative ecological impacts of barrier decommissioning are likely to become key considerations in the management of river infrastructure in non-perennial systems in drying climatic regions.

Given the complex impacts of instream barriers, both positive and negative, sound socioeconomic and ecological information should underpin decisions surrounding barrier decommissioning. The current study proposes specific criteria pertaining to fishes that merit consideration in barrier removal decisions in drying climatic regions with non- perennial rivers, and addresses the implications of climate change for these criteria. These include: ecological benefits of retaining barriers (e.g. providing artificial refuges); the spatial distribution of other barriers and key habitats (e.g. natural refuges) within watersheds; the presence of existing fish passage infrastructure; and biological characteristics of native fishes (e.g. resilience to climate change).

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To this end, a process for identifying, assessing and prioritising instream barriers for consideration of removal or remediation that incorporated these criteria, was developed and trialled in six catchments in south-western Australia. A total of 64 potentially significant barriers to fish migration were identified and prioritised according to their potential impact on fish migration. The highest ranked barriers tended to be located in the lower reaches of catchments, impacted higher proportions of resident fish assemblages, blocked access to higher proportions of available habitat (including refuges), and were less frequently drowned-out and passable to fish. Low-altitude aerial surveys (using helicopter) proved highly effective for validating information on artificial instream barriers gathered during a desktop review. The study also demonstrated that aerial surveys are especially useful in mapping permanent refuge pools and natural barriers, which are crucial to understanding distributions and movements of fishes in intermittent systems, but for which data are largely non-existent in this region. Gathering these data would have proved extremely difficult without the use of helicopter in these remote and largely inaccessible (by ground) catchments.

This study undertook the first field application of Passive Inducer Transponder

(PIT) technology to study the movement patterns of freshwater fishes in Western

Australia. Mature G. truttaceus were tagged and their movement patterns through a vertical slot fishway elucidated during a 5 month period that encompassed the annual breeding season. A total of 25 of the 144 (~16%) PIT-tagged individuals were detected at least once during the study. Sixteen individuals successfully passed through the fishway at least once, with some individuals ascending and descending on multiple occasions.

Successful passage through the fishway tended to coincide with flow pulses, although no passage events were recorded during a significant flood pulse that inundated the weir and fishway. The lag period between consecutive passage events for individual fish ranged viii from 1 to 116 d. Mean duration of fishway ascent (57.44 ± 11.50 min) was longer than the mean duration of descent (22.70 ± 3.04 min) and fish were detected in the fishway more frequently during the night (74.6% of total detections) compared to the day

(25.4%).

The number of PIT-tagged fish ascending the fishway was unexpectedly low, suggesting that spawning sites exist downstream of the weir, thus precluding the need for mature fish to undertake an upstream potamodromous migration via the fishway. A proportion of the nine fish that were detected at the fishway but did not pass through, visited the fishway entrance regularly over a period of days to weeks. It is unlikely that these fish were physically incapable of passaging the fishway, as the structure was specifically designed for this purpose; rather, this observation may represent the use of the fishway as habitat and could be interpreted as evidence of home-ranging behaviour in this species. Possible reasons for these observations are discussed.

Widespread sampling undertaken during the study extended the known range of one of the region’s rarest teleosts, N. pygmaea, a species that was previously known only from a small section of the Hay River. The study revealed that it occupies parts of the

Denmark and Kent rivers, and an outlying population was also discovered in Lake Smith

~200 km west of the nearest known population in the Kent. Its extent of occurrence is

3,420 km2 and area of occupancy is only 10 km2.

This study elucidated aspects of the biology of N. pygmaea in the Hay River system with the aim of determining key threats to its persistence. The population was found to undertake a short potamodromous migration into the seasonally flowing

Mitchell River to spawn in August and September before retreating to permanent refuge pools in the Hay River mainstem as flows abated in spring. It is a serial spawner and ix reached maturity at the end of its second year of life. No spawning females were captured in the salinised Hay River main channel despite extensive sampling, suggesting that access to non-salinised habitat is vital for reproduction in this population. The biology was compared with two sympatric percichthyids and discussed in reference to the considerable disparity in population size that was estimated in a baseflow refuge pool in the Hay River using mark-release-recapture methods. This was the first time the abundances of native freshwater fishes had been determined in a natural river system in south-western Australia.

The highly restricted area of occupancy of this species was interpreted as being indicative of a recent historical decline within the rivers that it inhabits, probably due to a limited salinity tolerance. The species is considered susceptible to catastrophic losses from potential perturbations such as prolonged drought, drying of critical baseflow refuges, increasing salinisation, and introductions of alien species. A number of refuge habitats for this species were identified throughout its restricted range, including two artificial refuges (fire-management ponds) in the Denmark catchment. There is potential to mitigate the threat of population declines or losses of this species by providing additional artificial refuge habitats, and further research into the efficacy of such a conservation strategy should be prioritised.

The information gathered during the study on the distribution, life history, threats and population size for N. pygmaea were used to assess its conservation status. It qualified as Endangered under IUCN Red List assessment criteria, and was recently added to Schedule 1 of Western Australia’s Wildlife Conservation Act 1950.

The current study is the first to develop a barrier prioritisation process that specifically accounts for the influence of climate change on the impacts of instream x barriers on fish migrations, which should enhance the robustness of decisions surrounding their removal or remediation in order to enhance fish migration. Climate change will become increasingly important in management decisions due to its direct and indirect impacts on fishes, particularly the effect it will have in exacerbating the impacts of instream barriers and water abstraction, and in reducing the quality and quantity of refuge habitats. The knowledge gained here on the life history and stressors impacting two of the continent’s rarest and most endangered fishes was incorporated into the barrier prioritisation process and will additionally assist environmental managers with the task of conserving the region’s aquatic biodiversity in the face of escalating threats.

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Table of Contents

Declaration ...... i

Acknowledgements ...... ii

Abstract ...... v

Chapter 1

General Introduction ...... 1

1.1 Freshwater fishes of south-western Australia ...... 1

1.2 Habitat degradation and secondary salinisation ...... 6

1.3 Alien species introductions ...... 7

1.4 River regulation and water abstraction ...... 9

1.5 Anthropogenic climate change ...... 13

1.6 Aims and objectives of the study ...... 17

Chapter 2

Implications of climate change for mitigating the impacts of instream barriers on migratory fishes

2.1 Introduction ...... 18

2.2 Restoring fish passage: barrier modification versus decommissioning ... 21

2.3 Ecological consequences of barrier decommissioning ...... 22

2.3.1 Potential positive impacts ...... 22

2.3.2 Potential negative impacts ...... 23

2.4 Implications of climate change for the decommissioning of instream barriers ...... 24

2.4.1 Impacts of climate change on freshwater fishes ...... 24

2.4.2 Exacerbation of negative impacts of instream barriers due to climate change ...... 26

2.4.3 Enhancement of the ecological value of barriers due to climate change ...... 28

2.4.4 Influence of climate change on the socioeconomic factors of barrier decommissioning ...... 29 xii

2.4.5 The impacts of barrier decommissioning on fishes and the influence of climate change ...... 31

2.5 Conclusion ...... 34

Chapter 3

A novel method for identifying fish migration barriers and prioritising their removal in drying climatic regions

3.1 Introduction ...... 35

3.1.1 Prioritisation methods in practice ...... 37

3.1.2 Considering synergistic impacts of climate change and instream barriers in prioritisation processes: south-western Australia as a model region ...... 39

3.2 A barrier identification and prioritisation process for drying climatic regions ...... 41

3.2.1 Project planning and development ...... 41

3.2.2 Validation of barrier information ...... 45

3.2.3 Barrier prioritisation ...... 46

3.2.4 On-ground works and monitoring ...... 50

3.2.5 Project extension ...... 52

3.3 Pilot study ...... 52

3.3.1 General methods ...... 53

3.3.2 Aerial surveys ...... 54

3.3.3 Prioritisation process ...... 56

3.4 Results ...... 57

3.4.1 Barrier prioritisation ...... 63

3.5 Discussion ...... 67

3.5.1 Aerial survey methods ...... 68

3.5.2 Stakeholder engagement ...... 69

3.5.3 Prioritisation of instream barriers ...... 71

3.6 Conclusion ...... 73 xiii

Chapter 4

Use of a vertical-slot fishway by a critically endangered migratory freshwater fish

4.1 Introduction ...... 76

4.2 Methods ...... 81

4.2.1 Site description ...... 81

4.2.2 Fishway ...... 84

4.2.3 PIT tags and detection system ...... 84

4.2.4 Population size estimation ...... 87

4.2.5 Environmental variables ...... 88

4.3 Results ...... 89

4.3.1 Detections of PIT-tagged fish ...... 89

4.3.2 Fishway movement patterns ...... 89

4.3.3 Influence of environmental variables on fishway movements .... 92

4.3.4 Population size estimate ...... 95

4.4 Discussion ...... 97

4.4.1 Fishway passage and flow ...... 97

4.4.2 Diel trends in fishway utilisation ...... 98

4.4.3 Site fidelity and fishway residency ...... 99

4.4.4 Limitations of the study ...... 101

4.4.5 Management implications ...... 103

Chapter 5

The biology and migration patterns of the Little Pygmy Perch, Nannoperca pygmaea (Teleostei: Percichthyidae), including comparisons with two sympatric percichthyids

5.1 Introduction ...... 107

5.2 Methods ...... 111 xiv

5.2.1 Study area ...... 111

5.2.2 Distributional sampling and fish monitoring ...... 114

5.2.3 Biological characteristics ...... 116

5.2.4 Environmental variables ...... 117

5.2.5 Visible Implant Elastomer (VIE) tagging and population size estimation ...... 118

5.3 Results ...... 122

5.3.1 Distribution of N. pygmaea ...... 122

5.3.2 Biological characteristics of N. pygmaea ...... 124

5.3.3 Migrations of pygmy perches ...... 132

5.3.4 Environmental variables ...... 136

5.3.5 Population size estimation in Hay River baseflow refuges ...... 140

5.4 Discussion ...... 145

5.4.1 Distribution of N. pygmaea ...... 145

5.4.2 Biology and migration ...... 146

5.4.3 Population estimates ...... 152

5.4.4 Management implications and future research requirements ...... 153

Chapter 6

General conclusions ...... 161

6.1 Barriers to fish migration in a drying climatic region ...... 162

6.2 Fishway use by critically endangered G. truttaceus during its breeding period ...... 164

6.3 Life history of the Little Pygmy Perch ...... 166

6.4 Interactions between climate change and key anthropogenic stressors of freshwater fishes in south-western Australia ...... 168

6.5 Recommended topics for further investigation ...... 170

6.6 Closing remarks ...... 174 xv

References ...... 176

Appendix 1 ...... 215

Appendix 2 ...... 218

1

Chapter 1

General Introduction

1.1 Freshwater fishes of south-western Australia

The south-west of Western Australia (Figure 1.1) is regarded as a globally significant centre of biodiversity exemplified by its exceedingly speciose flora (Olson & Dinerstein

1998; Myers et al. 2000). In contrast, the region’s indigenous freshwater teleost fauna is extremely depauperate, comprising only 11 described species; however, the percentage of endemic species (~82%) is the highest of any Australian ichthyological province (Table

1.1; Morgan et al. 2011; Unmack 2013). The inclusion of three additional ‘cryptic’ percichthyid species, currently awaiting formal description (see Unmack et al. 2011;

Unmack 2013), further increases this percentage.

A meta-analysis of factors explaining fish species richness in drainage basins around the world demonstrated that size (i.e. area) and annual discharge are highly correlated with species richness (Oberdorff et al. 1995). The paucity of fish species in south-western Australia is, therefore, hardly surprising, given the influence of the prevailing Mediterranean climate on streamflow throughout the region. Many rivers in south-western Australia are intermittent, contracting to a series of disconnected permanent pools during the hot, dry summer (Morrongiello et al. 2011). While winter rainfall is substantial near the coast, it declines sharply inland (BoM 2015), thus restricting the size and mean annual discharge of the region’s river basins (Allen 1982;

Berra 1998). 2

Figure 1.1. Map of south-western Australia showing major cities and regional towns,

topographic features, and the predominantly deforested agricultural area

known as the ‘wheatbelt’ (delineated by a yellow boundary).

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Table 1.1. Native freshwater fishes of south-western Australia, showing endemicity (to south-western Australia); migration category (An = anadromous, P = potamodromous, Am = amphidromous, L = local movements); and conservation status (CE = critically endangered). N.B. EPBC 1999 = Environment Protection and Biodiversity Conservation Act 1999; WCA 1950 = Wildlife Conservation Act 1950; ASFB = Australian Society for Fish Biology. Species with roman numerals (e.g. Bostockia porosa I) indicate cryptic species that have not yet been scientifically described (see Unmack et al. 2011; Unmack 2013)

Conservation status

EPBC Species name Distribution Migration WCA 1950 ASFB 1999 Geotridae Geotria australis non-endemic An - - - Plotosidae ‘Tandanus’ bostocki endemic P - - - Galaxias occidentalis endemic P - - - Galaxias maculatus non-endemic Am/P - - - * Galaxias truttaceus non-endemic Am/P CE Schedule 2 CE Galaxiella munda endemic P/L - Schedule 3 - Galaxiella nigrostriata endemic L - - - Lepidogalaxiidae Lepidogalaxias salamandroides endemic L - - - Percichthyidae Bostockia porosa endemic P - - - Bostockia porosa II endemic P - - - Nannatherina balstoni endemic P/L V Schedule 3 V Nannoperca vittata endemic P - - - Nannoperca vittata II endemic P - - - Nannoperca vittata III endemic P - - - Nannoperca pygmaea endemic P - Schedule 2 CE

* conservation status applies only to the Western Australian populations of this taxon.

4

Habitat heterogeneity and energy availability (i.e. primary productivity) are also excellent predictors of fish species richness in river basins worldwide (Guégan et al.

1998), supporting classical ecological theories pertaining to species-area (Preston 1962;

MacArthur & Wilson 1963, 1967) and species-energy (Wright 1983) relationships. These theories are underpinned by the assumption that greater numbers of species are able to co-exist in larger areas or those with higher primary productivity (i.e. energy) due to the increased diversity of available habitat and food resources, hence ecological niches, that such areas provide (MacArthur & Wilson 1967; Connor & McCoy 1979; Wright 1983;

Williamson et al. 1988). In south-western Australia, billions of years of weathering of the ancient landmass upon which it is formed and a prolonged period of tectonic stability have shaped a relatively flat and featureless landscape (Hopper & Gioia 2004) that lacks aquatic habitat diversity. For example, altitudinal habitat gradients and deepwater lacustrine systems are non-existent, evidently due to a lack of mountain building, volcanism and/or rifting in the region’s geologic history. Additionally, primary productivity is very low in south-western Australian rivers (Bunn & Davies 1990; Storey et al. 1993; Pen 1999), which may further explain the region’s depauperate fish fauna.

The high prevalence of endemism in south-western Australia is mostly attributable to vicariance resulting from a prolonged period of isolation from the rest of

Australia that stretches back ~30 Ma when the interior became arid (Frakes 1999). The formation of the Nullarbor Plain (an uplifted shallow seabed) during the mid-Miocene

(~15 Ma) (Sheard & Smith 1995; Webb & James 2006) is also thought to have contributed to this isolation (Unmack et al. 2011; Unmack 2013). While some ancient lineages unique to the region are still extant (i.e. Lepidogalaxias salamandroides and

‘Tandanus’ bostocki), the remaining south-western endemics belong to two families (i.e.

Galaxiidae and Percichthyidae) and all have phylogenetic affinities with species in 5 temperate eastern Australian fresh waters (Unmack et al. 2011; Unmack 2013). Two additional freshwater teleosts (Galaxias maculatus and Galaxias truttaceus) and an agnathan (Geotria australis) are also shared between the two regions, indicating that population connectivity still exists, at least for these diadromous species (Johnston et al.

1987; Berra et al. 1996; P. Unmack, pers. comm.).

Some of these species have suffered dramatic range declines during the past century (Morgan et al. 2003; Galleotti et al. 2010; Beatty et al. 2014a; Morgan et al.

2014), principally due to habitat loss, whilst others appear to be naturally restricted to narrow ranges (Morgan et al. 1998, 2011, 2014; Unmack 2001; Allen et al. 2002).

Consequently, several of the most imperilled species are now legislatively acknowledged as threatened at both State (Wildlife Conservation Act 1950) and Federal (Environment

Protection and Biodiversity Conservation Act 1999) level, and these species are the focus of some of the chapters that follow. The conservation status of the entire regional aquatic fauna faces increasing jeopardy due to a number of anthropogenic stressors and environmental threats including secondary salinisation, agricultural activity, habitat dewatering, water abstraction, river regulation, alien species introduction, and climate change (Morgan et al. 2003; 2004b; Morrongiello et al. 2011; Beatty & Morgan 2013;

Beatty et al. 2014a; IPCC 2014; Hope et al. 2015). The principal threats implicated in past and projected future declines of indigenous fishes in south-western Australia are reviewed in more detail below. 6

1.2 Habitat degradation and secondary salinisation

Land clearing in south-western Australia has occurred on a massive scale since European colonisation, with ~65% of native vegetation cleared for agriculture (Hobbs & Mooney

1998). Most land has been cleared in lower rainfall zones in northern and eastern parts of the region (i.e. the ‘wheatbelt’; Figure 1.1), which has resulted in dryland salinity

(Malcolm 1982; Schofield et al. 1988; Mayer et al. 2005). The phenomenon is caused by reduced transpiration and increased groundwater recharge when deep-rooted native vegetation is replaced with shallow-rooted annual crops, resulting in the deposition of soil-stored salts near the surface as the water table rises (Wood 1924; Cryer 1986; Peck &

Williamson 1987; Williams 2001), and causing secondary salinisation of rivers draining these areas (Peck & Hatton 2003; Mayer et al. 2005). Alarmingly, more than 50% of annual flow in the region’s 30 largest river systems is now brackish or saline (Mayer et al. 2005).

Salinisation has profoundly impacted fish communities in rivers such as the

Blackwood (the region’s largest by area and mean annual discharge), with euryhaline species that typically occur in estuaries (e.g. Leptatherina wallacei, Pseudogobius olorum) colonising salt-affected habitats hundreds of kilometres inland, whilst stenohaline (i.e. salt-sensitive) species have contracted to forested segments fed by fresh groundwater inputs from the Yarragadee Aquifer (Morgan et al. 2003; Beatty et al. 2009,

2010, 2014a). Dryland salinity and secondary salinisation have also had significant socioeconomic impacts in reducing agricultural productivity and profitability, as well as reducing the availability of potable water resources, which are already scarce and overexploited in the region (Mayer et al. 2005). 7

Further degradation and loss of aquatic habitats has also occurred in other parts of the region, most notably on the Swan Coastal Plain (Figure 1.1) where human population densities are high. Extensive clearing of native vegetation and drainage modification for agricultural and urban land development has resulted in the loss of ~80% of wetland habitat on the Swan Coastal Plain (Davis & Froend 1999; Bekle & Gentilli 1993;

Brearley 2005). Most of the remaining wetlands lack native freshwater teleosts and are heavily degraded due to excessive nutrient input (i.e. eutrophication), pollution, salinisation, and the introduction of alien species (Davis & Froend 1999; Morgan et al.

2004; Hourston et al. 2014). These habitat losses are inferred to have caused significant range declines for three south-western endemic fishes (i.e. Galaxiella munda, Galaxiella nigrostriata, and Nannatherina balstoni). These species tend to only occur in relatively pristine habitats in the south-west corner of the State, but have outlying populations hundreds of kilometres to the north, with no populations in the intervening area. In the absence of distributional data predating European colonisation, it is hypothesised that these isolated northern populations are relicts, and that these species have been extirpated from the southern Swan Coastal Plain due to habitat loss and degradation (McLure &

Horwitz 2009; Galeotti et al. 2010; Morgan et al. 2011, 2014).

1.3 Alien species introductions

Regrettably, another characteristic of the freshwater fish fauna of south-western Australia is the prevalence of alien species (Morgan et al. 2004, 2011; Beatty & Morgan 2013).

The earliest fish introductions in south-western Australia (i.e. Perca fluviatilis,

Oncorhynchus mykiss, Salmo trutta) were Government sanctioned for the purpose of creating enhanced recreational angling opportunities (Coy 1979) and for the biological 8 control of mosquitos (i.e. Gambusia holbrooki) (Mees 1977). Since the 1970s however, there has been a 75% increase in the number of fishes introduced to the region, mostly of aquarium and aquaculture species (e.g. Carassius auratus, Cyprinus carpio, Geophagus brasiliensis), presumably due to intentional or unintentional release, or via escape from ponds during flooding (Morgan et al. 2004, 2011; Beatty & Morgan 2013). There are now more alien fish species in the region than indigenous fish species and this trend is likely to continue in the future (Beatty & Morgan 2013).

Invasive alien species have been shown to have adverse impacts on indigenous species through predation, competition for food and habitat resources, agonistic behaviour, habitat degradation, and as vectors of diseases and parasites (e.g. Fletcher et al. 1985; Pen & Potter 1992; Gill et al. 1999; Morgan et al. 2002, 2004; Morgan &

Beatty 2007; Tay et al. 2007; Lymbery et al. 2010). These impacts will increase as more species are introduced and as those that are already established disperse naturally and via anthropogenic means (Beatty & Morgan 2013; Lymbery et al. 2014).

Although there are some examples where populations of alien species have been extirpated using interventionary measures (e.g. Lintermans 2000; Allen et al. 2013), generally once established they are impossible to eradicate from the wild (Horwitz 1990;

Rowe et al. 2008). Whilst early detection through regular field monitoring and rapid response to control/eradicate incursions of alien species are sound management policies

(e.g. Ayres & Clunie 2010), the most effective means of mitigating the impacts of introduced species is prevention rather than cure. This is probably best achieved through educational programs designed to raise awareness of the negative impacts that these species have on aquatic ecosystems and native biodiversity (see for example DoF 2013).

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1.4 River regulation and water abstraction

River regulation refers to the modification of water levels or flow regimes by artificial means for various anthropogenic purposes including potable water supply, irrigation, hydropower, flood control, drainage, transport, and navigation (Craig & Kemper 1987;

Petts 1999). Rivers are most commonly regulated via the construction of artificial barriers such as dams and weirs. In south-western Australia, large dams have been constructed on the main-channels of most of the rivers draining the Darling Range (Figure 1.1) to supply human water demands in the region. Further south, the main channels of most major rivers systems are undammed (many due to secondary salinisation); however, many large dams have been built on non-salinised tributaries to supply potable water to rural communities. A substantial number of small dams, weirs and other instream barrier types have also been constructed throughout the region, but the exact number has not been previously reported in the literature.

Instream barriers can have numerous and far-reaching impacts on lotic systems and aquatic faunal communities (see Chapter 2 for a review of instream barrier impacts).

These impacts can be particularly severe on fishes that are biologically adapted to migrate within rivers, with reduced habitat connectivity preventing or limiting access to vital habitats such as breeding grounds, seasonal feeding areas, nurseries and refuges (Lucas &

Baras 2001; Wofford et al. 2005; Koehn and Crook 2013).

Many native teleost fishes of south-western Australia are migratory, with most exhibiting a potamodromous life-cycle (Table 1.1; Pen & Potter 1991a, b; Pen & Potter

1990; Morgan & Beatty 2006; Beatty et al. 2010, 2014a). Myers (1949) first used the term ‘potamodromous’ to refer to fishes that complete their entire life-cycle within fresh waters but migrate within these waters usually for the purpose of breeding. Diadromous 10 fishes, in contrast, migrate between inland and oceanic waters, with this group being further subdivisible into anadromous, catadromous or amphidromous species (Myers

1949; McDowall 1997). Anadromy involves migration of adults (or sub-adults) from the sea to rivers for breeding, whereas catadromy involves a migration for the same purpose, but in the opposite direction. Geotria australis is the only species with an anadromous life-cycle that occurs in south-western Australia (Potter et al. 1979); no catadromous species are present. Amphidromy was a term developed by Myers (1949) for diadromous fishes “whose migration from fresh water to the sea, or vice-versa, is not for the purpose of breeding but occurs regularly at some other definite stage of the life-cycle”. McDowall

(1997, 2007) refined this definition to include only those fish that migrate immediately to sea as newly hatched larvae where they feed and develop for a period of a few weeks to months before returning to fresh water as juveniles. Two amphidromous species (i.e.

Galaxias maculatus and Galaxias truttaceus) are known from fresh waters of the south coast, although some populations are potamodromous (Morgan 2003; Chapman et al.

2006; Morgan & Beatty 2006; see also Chapter 4). Other species, such as Lepidogalaxias salamandroides, Galaxiella nigrostriata and some populations of Nannatherina balstoni undertake short, lateral movements to breed in flooded riparian vegetation (Pen et al.

1993; Morgan et al. 1995, 2000), but according to Myers (1949) such movements are not considered potamodromous, and are best described as local migrations.

Large dams completely cut off access to habitats for migratory aquatic fauna, whereas smaller barriers may become passable during periods of inundation or ‘drown- out’. Some fishes (e.g. non-native salmonids) are able to leap out of the water to bypass some barriers (e.g. Stuart 1962; Reiser & Peacock 1985), but whilst there is only limited published information on the leaping ability of south-western Australian native fishes

(Beatty et al. 2014b; Close et al. 2014), field observations over many years suggest that 11 barriers with a differential in head and tail water levels (i.e. head loss) ≥ 0.3 m are effectively impassable to most species from this region (S. Beatty & D. Morgan, pers. comm.). However, both G. australis and G. truttaceus provide exceptions, as they are able to ‘climb’ damp rock or concrete surfaces in order to bypass small dams and weirs during migrations (Morgan et al. 1998; Close et al. 2014), although this behaviour may be limited to juveniles in the latter species. Regardless, instream barriers reduce the proportion of fish able to bypass these migration obstacles as well as potentially influencing their food supply by impacting macroinvertebrate communities (Bunn &

Arthington 2002), and altering natural hydrological regimes, which can provide essential cues for spawning in some species (e.g. Morgan 2003). For example, the capture of early season flows in regulated streams can delay access to spawning sites and potentially shorten the duration of the breeding season, thus impacting recruitment for some species

(sensu Beatty et al. 2014a). In catchments with large dams (hence, large volumes of impounded water), the hydrological impacts of water abstraction can be regulated through allocation of ‘environmental flows’ (i.e. water released from dam) in order to maintain ecological function in habitats downstream of dams and mimic the natural flow regime, albeit at reduced levels (e.g. Radin et al. 2007; Donohue et al. 2009). However, this is not the case for the multitude of smaller dams, weirs and other instream barriers throughout the region.

The impacts of barriers on fish migration can be mitigated through their modification or removal, and a commonly used strategy to achieve this is the installation of fish passages (also known as fish ladders or fishways) (Katopodis 1992; Odeh 1999; see also Chapters 2 & 3). However, only seven fishways have been constructed in the region, and these only provide passage for a portion of the migratory freshwater fauna

(i.e. Galaxias spp., Geotria australis, and limited movements of N. vittata) (e.g. Morgan 12

& Beatty 2006; Beatty et al. 2007, 2014b). A seemingly simpler and more effective solution for mitigating the impacts of barriers on fish migration is decommissioning or removal; however, these structures are rarely redundant and can sometimes offer ecological benefits such as the provision of aquatic refuges in their associated reservoirs, which is an ecological function that may become increasingly valuable due to climate change (see section 1.5; Chapters 2-3), or by limiting the dispersal of invasive alien species (see also Chapters 2-3).

Abstraction of groundwater can also have hydrological impacts in both lotic and lentic systems (CSIRO 2009b; Petrone et al. 2010). Groundwater can be vital in maintaining ecological functionality in a range of aquatic habitat types including wetlands, perennial lotic systems, and permanent refuges in intermittent streams (Balla &

Davis 1993; Beatty et al. 2010, 2014a). Anthropogenic water abstraction can take other forms including pumping of river water, non-compliance by private landholders in the installation and use of flow-bypass valves in dams, and the establishment of timber plantations in former wetland areas which can cause draw down of water tables and threaten ephemeral aquatic habitats (DoW 2009). Water abstraction (both surface and subsurface) is already very close to the maximum sustainable level necessary to maintain aquatic ecological function in some of the region’s catchments (e.g. Donohue et al. 2010;

Green et al. 2011) and water license allocations and environmental flow releases will need to be closely monitored and managed adaptively in a drier future.

13

1.5 Anthropogenic climate change

Strong consensus now exists among the world scientific community that climatic changes observed since the 1950s are almost certainly attributable to human actions, principal among which is the emission of ‘greenhouse gases’ (IPCC 2014). Global Climatic

Circulation Models (GCMs) project a mean global temperature rise of 1-7°C by the end of the 21st century (IPCC 2013), and the resultant physico-chemical and hydrological alterations to freshwater environments are likely to have significant impacts on freshwater fishes worldwide (Ficke et al. 2007; Morrongiello et al. 2011).

The impacts of a changing global climate have already been felt in south-western

Australia, where mean annual rainfall and surface runoff have declined by 10-15% and

~50%, respectively, since the mid-1970s (Figure 1.2; Suppiah et al. 2007; Petrone et al.

2010; Silberstein et al. 2012). Declines have been most pronounced in the May-July period (i.e. early winter), as rain-bearing weather systems have reduced substantially in frequency and intensity (Hope et al. 2006; Silberstein et al. 2012). The disproportionately large reduction in surface runoff cf. rainfall (a ratio of ~3:1) is most likely attributable to coincidental declines in soil moisture and groundwater levels across the region (CSIRO

2009b), which evidently play a crucial role in determining the conversion of rainfall to surface runoff (Croton & Reed 2007; Petrone et al. 2010).

The impacts of anthopogenic climate change on the region’s freshwater fish fauna are obscured by patchy historical data on ecology and distribution (Beatty et al. 2010;

Morrongiello et al. 2011); however, recent evidence has emerged linking climate change to the decline of two species (Ogston et al. 2016), i.e. Lepidogalaxias salamandroides and Galaxiella nigrostriata. These fishes inhabit ephemeral pools and wetlands in the 14

1000

900

800

700

600

500

400 Runoff (GL) 338

300

200 177

100 78

0 1910 1920 1930 1940 1950 1960 1970 1980 1990 2000 2010 Year

Figure 1.2. Annual surface flow into reservoirs that supply the city of Perth, Western

Australia, between 1911 and 2014. Numbers indicate mean runoff over the

time periods indicated by horizontal lines (data source Water Corporation,

Western Australia).

15 south-western corner of the State and are uniquely adapted to survive habitat desiccation by burying themselves in the moist substrate and aestivating until winter rains refill their habitat (Allen & Berra 1989; Berra & Allen 1989; Pusey 1990; Morgan et al. 2000).

Recent sampling revealed that these species were absent from ~30% of sites where they had been collected in the 1980s and 1990s, and groundwater declines due to climate change were inferred to be responsible for this result (Ogston et al. 2016).

Dramatic recent declines in the abundance of Galaxiella munda in the Margaret

River system have also been ascribed to climate change (Allen et al. 2015; DoW 2015), although groundwater abstraction from the Yarragadee Aquifer may have also contributed. Galaxiella and Lepidogalaxias are typically short-lived (life-span of 1-5 years) (Pen et al. 1991, 1993; Morgan et al. 2000), rendering them vulnerable to population crashes during sustained periods of extreme environmental conditions that reduce breeding success and recruitment rates (Allen et al. 2015; Ogston et al. 2016). The declines of these species may indeed serve as a bellwether of the potential for climate change to impact other aquatic fauna in the region.

Global Climatic Models unanimously project that by 2030, rainfall and surface water runoff in south-western Australia will decline by a further ~8% and ~24%, respectively, and the number of no-flow days in lotic systems will increase by up to 4 months (Suppiah et al. 2007; CSIRO 2009a, b; Barron et al. 2012; Silberstein et al.

2012). Warming of around 0.5-1.1 °C is projected by 2030 with drought periods and days of extreme heat also projected to increase (Hope et al. 2015) leading to higher evaporation rates (CSIRO 2009a, b; Hope et al. 2015). Further declines in groundwater levels (>10 m in some parts of the region) are also projected (Barron et al. 2012). Climate change will not only have hydrological impacts that will directly affect aquatic 16 ecosystems, but they will also exacerbate the impacts of other key threatening processes

(see section 6.3).

Significant fish biodiversity losses have been forecast, particularly in global regions with semi-arid or Mediterranean climates (Tedesco et al. 2013) due to streamflow reductions resulting from climate change and increasing levels of water abstraction

(Xenopoulos et al. 2005; Xenopoulos & Lodge 2006). Western Australia’s region, southern America (i.e. Chile and Argentina) and southern parts of middle-eastern

Asia (i.e. Iran, Afghanistan, and Saudi Arabia) are projected to suffer the highest numbers of extinctions of aquatic species (Tedesco et al. 2013). Although no specific projections have been made for south-western Australia, climate change has already had a profound hydrological impact in some parts of the region. For example, numerous streams in the

Darling Range that were perennial prior to the climatic shift that occurred in the 1970s now have intermittent flow (Petrone et al. 2010). The impact that continued warming and drying will have on freshwater ecosystems in the region that are already impacted by numerous anthropogenic stressors is likely to be significant (Morrongiello et al. 2011).

There are serious concerns about the implications of climate change on fishes that have behavioural and life-history adaptations that are finely attuned to natural flow regimes

(Morgan et al. 2011; Morrongiello et al. 2011). Further declines in rainfall, surface flow and groundwater have the potential to decrease the annual duration of seasonal flow or inundation of key spawning and nursery habitats, as well as the quality and quantity of permanent refuge habitat (e.g. Beatty et al. 2014a; Morgan et al. 2014; Ogston et al.

2016; see also Chapters 2, 3 & 5). The anticipated declines in breeding and recruitment of freshwater fishes (Morrongiello et al. 2011) and reduction in the biological carrying capacity of freshwater ecosystems due to anthropogenic climate change may lead to further localised extirpations or even extinctions (IPCC 2014; DEC 2008). 17

1.6 Aims and objectives of the study

The overarching aim of the thesis was to investigate how a drying climate may influence the impacts of instream barriers on freshwater fish migration, using south-western

Australia as a model region due to its recent and projected severely drying climate, its numerous instream barriers, and its highly migratory freshwater fish fauna. To achieve this, the major aims of the study were:

 To review existing information on the impacts of instream barriers, including how

they are mitigated, the global trends in their construction and

removal/remediation, the factors underpinning approaches that have been used to

prioritise their mitigation, and the implications of projected climate change.

 To develop a process for identifying and prioritising instream barriers for

removal/remediation that is applicable to drying climatic regions such as south-

western Australia.

 To apply the newly developed prioritisation process across a number of case

study catchments.

 To elucidate important biological traits, migration patterns, population sizes,

distributional range, and key threats for two of the region’s most endangered fish

taxa in order to increase the rigour of the pilot application of the barrier

prioritisation process.

 And, to make specific recommendations regarding the management of the

region’s most endangered freshwater fishes in the face of climate change and

other key anthropogenic stressors such as instream barriers. 18

Chapter 2

Implications of climate change for mitigating the impacts of instream

barriers on migratory fishes

2.1 Introduction

Freshwater ecosystems are among the most endangered in the world due to the combined effects of water abstraction, habitat destruction, flow modification, pollution and invasive species (Dudgeon et al. 2006). These ecosystems are also likely to be impacted by climate change in the future because their isolation and fragmentation within terrestrial landscapes severely limits the dispersal ability of freshwater biota (Millennium

Ecosystem Assessment 2005; Woodward et al. 2010). Climate change may affect freshwater biota directly, such as through the impact of increasing water temperature on physiological processes (Pankhurst & Munday 2011), indirectly, such as through the impact of altered flow regimes on ecological processes (Tisseuil et al. 2012), or by more complex, tertiary impacts such as interactions with river regulation (sensu Koehn et al.

2011).

Over half of the world’s large river systems (i.e. with a mean annual discharge ≥

350 m3s-1 at any point in the catchment prior to human regulation) are now affected by dams (Nilsson et al. 2005). Service (2011) reported that in the United States alone there were ~2,000,000 total river blockages. Globally, large dam (i.e. crest height > 10m) construction increased rapidly in the post-World War II period and peaked in the 1970s

(WCD 2000). The majority of these structures are located in China and India, where dam construction has continued more rapidly in comparison with the rest of the world (WCD 19

2000). There is no global estimate of the total number of small instream barriers (i.e. crest height < 10 m) (see also Lejon et al. 2009).

Whilst dams and other constructed instream barriers provide many benefits to humans, they also have numerous detrimental effects on aquatic ecosystems (e.g. Ward &

Stanford 1979; Poff et al. 1997; Bunn & Arthington 2002). Abiotic impacts include habitat fragmentation, reductions in habitat quality and complexity, and disruption to processes of erosion, sediment transport, channel scouring and nutrient cycling (e.g. Junk et al. 1989; Kondolf 1997; Górski et al. 2012). The biological responses to these impacts can include shifts in community composition, loss of species richness and abundance, and distributional range contractions of native species (e.g. Bain et al.1988; Sheer & Steel

2006). In many cases the impacts of dams are greater at the regional scale than at the local scale, because of the importance of tributary-main channel and upstream- downstream connections in a drainage basin (Van Looy et al. 2014). The significance of the impacts that instream barriers have on aquatic ecosystems is reflected in their recognition as a key threatening process that is acknowledged and integrated into legislative frameworks for conservation in jurisdictions such as Europe (Kemp &

O’Hanley 2010) and Australia (e.g. EPBC 1999).

The impacts of instream barriers on fishes include, inter alia, disruption of migratory pathways, creation of unfavourable habitats for native species, loss of riparian habitat through the inundation of river valleys upstream of barriers, alteration of natural flow regimes that provide cues for migration and spawning, and sedimentation of habitats upstream of barriers (e.g. Lucas & Baras 2001; Bunn & Arthington 2002; Katopodis &

Aadland 2006). Whilst there has been considerable research on mitigating the impacts of barriers on fish communities via the construction of fishways (e.g. Katopodis 1992;

Roscoe & Hinch 2010), attention has only recently been given to the efficacy of 20 decommissioning (i.e. removing) instream barriers (e.g. Bednarek 2001; Gregory et al.

2002; Feld et al. 2011).

The influence of climate change on the impacts of instream barriers and strategies for mitigating those impacts has received limited attention. This is a particular concern in regions where the relative ecological efficacy and cost-effectiveness of removing instream barriers versus the construction of fishways may be considerably influenced by climatic shifts. Many temperate regions, for example, have highly seasonal rainfall and flow regimes to which native aquatic fauna are adapted (Matono et al. 2012; Beatty et al.

2014a). Climate change may alter the impacts of instream barriers on aquatic ecosystems in these regions (Ficke et al. 2007; Cañedo-Argüelles et al. 2013), potentially leading to declines or extinctions of some freshwater fish species in regions transitioning to a drier climate (Davies 2010; Morrongiello et al. 2011; Beatty et al. 2014a). Therefore, there is an urgent need to better understand how climate change will influence the impact of instream barriers, and moreover, how it may influence the efficacy of strategies to mitigate those impacts.

This review aims to synthesise information that can assist in the development of rigorous processes for assessing the impacts of barriers and determining appropriate impact mitigation strategies, with the purpose of providing maximum benefit for migratory fish. The review examines the efficacy of retrofitting existing barriers with fishways versus decommissioning to mitigate their impacts. It then identifies the potential positive and negative impacts of decommissioning and explores the implications of climate change on the efficacy of barrier removal in drying climates, using case examples from south-western Australia that are applicable to other parts of the world experiencing climatic drying. 21

2.2 Restoring fish passage: barrier modification versus decommissioning

The dispersal of many fish species is impeded by instream barriers (e.g. Sheer & Steel

2006; Koehn & Crook 2013), and mitigation of this impact may be achieved via decommissioning (i.e. removal or breaching), barrier modification (e.g. retrofitting with fishways, fish locks or fish lifts), or by employing trap-and-haul methods (Gregory et al.

2002; Travade & Larinier 2002; Kemp & O’Hanley 2010). Because of the socioeconomic value of barriers (Gregory et al. 2002), modification has been more common than decommissioning. Globally, fishways have proven most effective at providing passage for large-bodied salmonids, but are much less effective for smaller-bodied, slower swimming species (Katopodis & Aadland 2006; Mallen-Cooper & Brand 2007; Noonan et al. 2012).

Complete physical removal of instream barriers circumvents inefficiencies of retrofitting barriers with fish passage facilities and is an effective means of increasing longitudinal connectivity in regulated rivers for migratory fishes. Instream barrier removal has increased markedly throughout the world since 1990 (e.g. Bednarek 2001;

Service 2011). The vast majority of large dams in the United States are approaching their life expectancy and dam decommissioning (averaging just under 60 removals per year over the past decade) now outpaces construction in that country (Bednarek 2001; Pohl

2003; American Rivers 2016). The situation in Europe is similar, although the rate of decommissioning is not as high (Lejon et al. 2009; European Rivers Network 2013). In

Australia, fishway construction has gathered pace, but less than 50 barrier removals

(mostly small structures) have been documented (see Nelson 2003; I&I NSW 2009;

O’Brien et al. 2010; State Water 2012). Reports of instream barrier removals from elsewhere in the world are scarce. 22

Barrier decommissioning can potentially provide substantial environmental benefits; however, it is rarely undertaken solely for environmental purposes. The main reasons for removing barriers include: 1) unacceptably high levels of impact (particularly socioeconomic impacts); 2) high maintenance costs; 3) expiration of the barrier’s useful lifespan (i.e. redundancy); or 4) a combination of the preceding reasons backed by law or policy (Stanley & Doyle 2003; Lejon et al. 2009). Economics is the usual driver of barrier removals, as the costs to repair aging infrastructure and/or install fish passes to meet safety and environmental standards usually greatly outweigh removal costs

(Sarakinos & Johnson 2003; Katopodis & Aadland 2006).

2.3 Ecological consequences of barrier decommissioning

Barrier decommissioning decisions involve trade-offs between positive and negative ecological and socioeconomic impacts (Bednarek 2001; Stanley & Doyle 2003;

McLaughlin et al. 2013). Such decisions involve risks and these need to be managed in an objective, structured fashion (McLaughlin et al. 2013). Opposition to proposed barrier decommissioning projects from stakeholders (including the general public) can be strident, and often stems from a lack of understanding of the function and value of rivers and dams (Sarakinos & Johnson 2003). This commonly results in the cancellation of projects or a compromised restorative approach being taken (e.g. Lejon et al. 2009).

2.3.1 Potential positive impacts

Artificial lentic water bodies are often hot spots for alien species (Bunn & Arthington

2002; Johnson et al. 2008; Rahel & Olden 2008) and barrier decommissioning can reduce 23 favourable habitat for introduced species by converting artificial lentic habitat back into its original lotic state. The restoration of more natural regimes of temperature, carbon/nutrient cycling, and sediment transport can also contribute to increased species richness, abundance, and biomass of fishes at formerly impacted sites (Hill et al. 1994;

Kanehl et al. 1997; Catalano et al. 2007). Reinstatement of longitudinal river connectivity following decommissioning can permit access to habitat beyond former barriers, leading to increases in recruitment and productivity within relatively short timeframes; particularly for anadromous salmonids (Pess et al. 2008). The improvement of lateral connectivity between rivers and floodplains following barrier removal can also benefit aquatic and dependent terrestrial fauna (Hill et al. 1994; Shuman 1995).

Fishes often sustain injuries while attempting to bypass barriers (e.g. swimming through hydropower turbines or over spillways) (Travnicheck et al. 1993; Coutant &

Whitney 2000). Such injuries have been estimated to result in the mortality of 5-20% of immature salmonids in some river systems impacted by hydroelectric dams in the north- western United States (Raymond 1979; Skalski 1998). Barrier decommissioning can mitigate these risks, as well as easing predation pressures on migratory species that congregate in unnaturally large numbers downstream of barriers (e.g. Buchanan et al.

1981; Ruggerone 1986).

2.3.2 Potential negative impacts

Unless carefully managed, mobilisation of accumulated sediments during barrier decommissioning can have severe impacts on habitats downstream through sediment deposition (which may contain toxins, heavy metals or nutrients) and erosion (Shuman

1995; Bednarek 2001; Stanley & Doyle 2003). The removal of barriers can also 24 potentially facilitate the spread of invasive species and exacerbate their impacts (Fausch et al. 2006; McLaughlin et al. 2007). There are many examples of instream barriers that limit the spread of such species, including some structures that have been intentionally built for this purpose (see McLaughlin et al. 2013).

By removing an instream barrier, the impounded water body immediately upstream is lost. Although such artificial lentic habitats can harbour alien fish species

(e.g. Rahel & Olden 2008), they have also been shown to provide important refuges for aquatic fauna in some non-permanent river systems (e.g. McNeil et al. 2011; Lintermans

2012). There is potential to increase the risk to native aquatic fauna through the loss of artificial refuges caused by barrier decommissioning if inadequate accessible refuge habitat exists elsewhere.

2.4 Implications of climate change for the decommissioning of instream barriers

2.4.1 Impacts of climate change on freshwater fishes

Global Climatic Circulation Models (GCMs) project a mean global temperature rise of 1-

7°C by the end of the 21st century (IPCC 2013). The resultant increase in water temperatures, decrease in dissolved oxygen, alterations to flow regimes, and changes in habitat quality and availability are likely to have a significant impact on freshwater fishes across the globe (Ficke et al. 2007; Morrongiello et al. 2011). The ecological responses to these abiotic impacts are likely to include community shifts, alteration of primary production and food webs, distributional range shifts and/or declines, and in some cases extinctions (Ficke et al. 2007; Morrongiello et al. 2011; Tedesco et al. 2013). Ecological changes in freshwater systems will largely by driven by alterations to physico-chemical 25 and hydrological regimes, and the disruption of habitat connectivity (Koehn et al. 2011;

Morrongiello et al. 2011; Beatty et al. 2014a). Species with broad physico-chemical tolerances (including some alien species) are likely to become dominant, whilst those with narrower tolerances or specialised habitat preferences are likely to decline in abundance or be extirpated from some regions (Davis et al. 2013).

A number of studies have considered the implications of climate change on fish distributions, mostly focussing on northern hemisphere temperate systems and cold-water species such as salmonids (see Comte et al. 2013). Although there are some exceptions

(e.g. Balcombe et al. 2011; Bond et al. 2011; Tedesco et al. 2013), most studies have concentrated on rising water temperature as the primary driver of change (e.g. Meisner

1990; Heino et al. 2009; Comte et al. 2013); however, in Mediterranean, semi-arid and arid regions, hydrological shifts are likely to have an equally, if not more, significant impact on fish communities (Jenkins et al. 2011; Morrongiello et al. 2011; Davis et al.

2013; Beatty et al. 2014a).

The lack of focus on hydrology cf. temperature as a driver of change in fish communities may be due to the uncertainty of rainfall projections in many global regions

(Tisseuil et al. 2012). However, a high degree of consistency in projected rainfall exists among climate change models for regions such as south-western Australia (IOCI 2008;

CSIRO 2009a; Hope et al. 2015). This region has already undergone a pronounced climatic shift since the mid-1970s, with declines in rainfall and surface discharge and increased temperatures, and GCMs unanimously project the drying trend to continue

(Suppiah et al. 2007; IOCI 2008; CSIRO 2009; Barron et al. 2012; Silberstein et al.

2012; see also section 1.5). South-western Australia therefore represents an ideal, if unfortunate, exemplar of the hydrological impacts of climatic drying and its implications for freshwater fishes (Beatty et al. 2014a). 26

The synergistic impacts of climate change and instream barriers will result in hydrological shifts that will test the resilience of freshwater fishes in south-western

Australia (Davies 2010; Koehn et al. 2011; Morrongiello et al. 2011; Beatty et al. 2014a).

Most species spawn and recruit in ephemeral streamlines and wetlands, with populations retreating to summer refuges (typically pools located in mainstems) where they persist during baseflow (Morgan & Beatty 2006; Beatty et al. 2014a). Spatiotemporal declines of these crucial seasonal habitats are likely under projected climate scenarios, as are declines in the hydrological connectivity between these habitats, and concomitant declines in the faunal carrying capacity of freshwater systems in the region seem highly probable (Beatty et al. 2014a). This is also likely to occur elsewhere in southern Australia and in other

Mediterranean and temperate climate regions where rainfall reductions are projected

(Morrongiello et al. 2011; Davis et al. 2013; Robson et al. 2013).

2.4.2 Exacerbation of negative impacts of instream barriers due to climate change

Hydrological and physico-chemical shifts driven by climate change may exacerbate some impacts of instream barriers on fishes (Hughes 2003). For example, declining flows are likely to shorten the temporal window available to migratory fishes for bypassing inundated barriers, which may lead to limited or failed recruitment. Increased water temperatures will cause an increase in the metabolic rate and oxygen demand for most fishes, influencing physiological processes such as locomotion (Schmidt-Nielsen 1990;

Kalff 2000). Furthermore, oxygen solubility is inversely related to water temperature

(Kalff 2000), therefore fish may become oxygen stressed in habitats lacking thermal refuges (Ficke et al. 2007). This will impact the swimming performance of fishes (e.g. 27

Brett 1971; Kutty & Sukumaran 1975) and may compromise their ability to ascend fishways or, in extreme cases of hypoxia, mortality may result.

The passage efficiency of fishways may also decline due to a mismatch between the temperature and flow patterns for which structures were originally designed and those projected with climate change (Kemp et al. 2011). For example, significant delays have been shown for salmonids to successfully pass a fishway at annual temperature maxima due to the impact of high temperatures on swimming performance (Caudill et al. 2013).

Climate change will exacerbate these effects, potentially creating thermal barriers at existing fishways unless modifications occur (Caudill et al. 2013).

The synergistic impacts of instream barriers and climate change also have the potential to exacerbate the impacts of invasive alien species. Rahel & Olden (2008) foreshadowed the increased likelihood of novel invasions by species that possess physiological thresholds matched to conditions that are likely to prevail under future climatic scenarios. Indeed, Beatty and Morgan (2013) found that the majority of introduced fishes already established in temperate south-western Australia were climatically mismatched, being native to tropical or sub-tropical regions. Moreover, global proliferation of dams and river infrastructure (e.g. inter-basin distribution networks) required to meet future human water demands has the potential to facilitate the further spread of invasive species (Waters et al. 2002; Rahel & Olden 2008). The beneficial role that barriers can play in restricting this spread may help offset the impacts of alien species (Rahel & Olden 2008; McLaughlin et al. 2013), and these contrasting roles will need to be duly considered when assessing barrier removal proposals.

The impoundment and abstraction of water can cause severe flow reductions in rivers, potentially leading to moisture stress in downstream ecosystems, and this impact 28 will undoubtedly be exacerbated by climate change (Hulme 2005; Palmer et al. 2008). In drying regions such as southern Australia’s Murray-Darling Basin, environmental flow releases from reservoirs have been used to counter moisture stress in floodplains that provide vital habitats for many fish species (Arthington & Pusey 2003). However, allocating adequate environmental water provisions from reservoirs to downstream ecosystems will be more challenging in drying climatic regions due to the diminished supply and increased socioeconomic demands for water (e.g. Pittock et al. 2013).

2.4.3 Enhancement of the ecological value of barriers due to climate change

Refuge habitats are essential for the persistence of aquatic organisms in non-perennial aquatic ecosystems. The potential loss of natural refuges under reduced rainfall and flow conditions in drying climatic regions may be offset to some extent by maintaining existing instream barriers and their associated impoundments, or by creating additional lentic habitats. Davis et al. (2013) identified a trade-off in arid zone refuge habitats between hydrological connectivity, which permits rapid dispersal, and climatic decoupling, which isolates the microclimate from the regional climate. In general, habitats with strong decoupling will act as evolutionary refuges, while those with high connectivity will provide ecological refuges. The artificial lentic habitats which result from instream barriers have the potential to act as both evolutionary refuges and (when coupled with fishways or other structures that allow fish passage) ecological refuges for freshwater fishes.

It is likely, therefore, that in many instances the ecological and conservation value of artificially created lentic refuges in drying climatic regions may increase in the future and this will need to be considered in barrier decommissioning proposals. For example, in 29

Australia’s Murray-Darling Basin, several artificial habitat types, including weir pools, have been identified as high priority refuges (McNeil et al. 2013). Whilst they were generally rated as having a lower ecological value than natural refuges, they were rated higher in terms of their likelihood “to persist during the most extreme or prolonged climatic disturbances”, therefore the ecological value of these artificial refuges was predicted to rise with increasing severity of drought (McNeil et al. 2013). The conservation value of such habitats may be further enhanced when coupled with interventionary actions such as the translocation of threatened fauna populations (e.g.

Hammer et al. 2013), riparian planting (Davies 2010), or provision of instream habitat

(e.g. large woody debris) to enhance the “naturalness” of artificial refuge habitats

(Erskine & Webb 2003).

2.4.4 Influence of climate change on the socioeconomic factors of barrier decommissioning

Climate change is projected to alter hydrological regimes globally (Palmer et al. 2008) and this will have major socioeconomic implications for the maintenance and management of river infrastructure (Pittock & Hartmann 2011). On the one hand, increasing human demands for water due to global population growth (Palmer et al.

2008; WWAP 2009; Pittock & Hartman 2011) will mean that the socioeconomic benefits delivered by instream barriers are likely to outweigh the environmental costs (Palmer et al. 2008). However, an increasing number of dams worldwide are also reaching the end of their lifespan and require imminent replacement, repair or removal (e.g. Bednarek

2001; American Rivers 2010; European Rivers Network 2013). Climate change may accelerate this process. For example, in regions where increased frequency of extreme 30 flooding is expected, dams will need to be reinforced and/or modified to mitigate the risk of overflow or structural failure (Pittock & Hartmann 2011). As the cost of maintaining aging river infrastructure typically far exceeds that of removal (e.g. Sarakinos & Johnson

2003), decommissioning is likely to become increasingly preferred on economic grounds.

Furthermore, in drying climatic regions, the utility and cost-effectiveness of maintaining storage dams in particular will reduce sharply as rainfall and surface flows decline, increasing the rate at which barriers lose economic viability. Therefore the rate of barrier decommissioning is likely to gather pace in these regions. Barrier decommissioning will also come under increasing consideration as a strategy to mitigate the hydrological impacts of climate change on lotic ecosystems that have intrinsic, rather than economic, value to humans, and on human populations directly by the elevated risk of catastrophic dam failure due to increased intensity and frequency of floods (WCD 2000; IPCC 2013;

Palmer et al. 2008; Pittock & Hartmann 2011).

Regions that experience intense climatic drying will eventually reach a tipping point where the benefits of many barriers will be outweighed by costs, leading to a potential scenario of extensive barrier removal. For example, Australia is currently facing pressure to increase its water supply in order to meet the demand of a growing population, with water consumption in major urban centres predicted to increase by 64-

107% over the next 50 years (SoE 2011). The relative contribution made by large storage dams to water supplies across southern Australia has declined in recent decades and this trend is projected to continue (Silberstein et al. 2012). For example, by 2060 large storage dams are anticipated to no longer provide a reliable contribution to the potable water supply in south-western Australia (Water Corporation 2009; SoE 2011). As a result, water security through diversity of supply is now the paradigm for this region, and is reflected by the increase in water conservation measures, and recycled and manufactured 31 water sources; particularly desalinisation (SoE 2011). Therefore, the combined effect of the finite lifespan of dams and their diminishing economic value as a reliable water source in regions that are transitioning to a drier climate (e.g. south-western Australia), is likely to increase the rate of dam obsolescence and decommissioning (Pittock &

Hartmann 2011).

2.4.5 The impacts of barrier decommissioning on fishes and the influence of climate change

It is vital that decisions on removal of instream barriers weigh the positive (e.g. increased stream connectivity; restoration of hydrological regimes) and potentially negative (e.g. loss of artificial baseflow refuges; potential spread of alien species) ecological impacts of removing instream barriers under both current and projected future climatic conditions. In order to determine whether or not the decommissioning of a particular barrier will result in a net ecological benefit, it is crucial that the hydrology and ecology of not only the artificial water body that would be impacted by the removal of the barrier, but also that of the entire watershed in which it is situated, is well understood, as are the probable impacts of projected climatic change on fluvial systems in the region.

In a hypothetical example of a stream network where the locations of barriers and spatial distribution and life-histories of its fishes are known, it might seem intuitive to remove a barrier to enhance seasonal access to a previously unreachable stretch of river for a migratory species. However, if the artificial impoundment created by this barrier is an important refuge for a threatened species, then it would not be prudent to decommission the barrier. Installation of a fishway may be an appropriate compromise in this situation, but if projected climate change is likely to result in the loss of the artificial 32 refuge, the decision to remove the barrier may be appropriate, although other management responses to conserve the threatened species would be necessary in this instance, such as creating artificial permanent refuge habitat for example.

This highly simplified example highlights the importance of sound ecological, hydrological and climatological knowledge of the watershed when assessing barrier decommissioning in relation to fishes. While it is generally recognised that the former two knowledge areas are vital in barrier removal decisions, the consideration of climate change in these processes has been virtually non-existent (see Chapter 3). Decisions surrounding the mitigation of instream barrier impacts involve a trade-off between the costs and benefits of removing versus retaining structures. Whilst socioeconomic considerations usually drive such decisions, ecological considerations are also important.

The following criteria pertaining to fishes warrant careful consideration in the decision making process, as do the ramifications of climate change on these criteria, which in many cases are likely to be substantial.

1) Ecological benefits of retaining barriers

Besides the obvious socioeconomic benefits, instream barriers can also provide ecological benefits (e.g. providing refuge habitat; limiting the spread of invasive species).

Some of these positive effects may become increasingly important in drying climates.

Arguably, the most amplified benefit of instream barriers and their associated impoundments will be their role in providing artificial refuges in regions where natural refuges decline in quality and quantity. Therefore, as part of instream barrier decommissioning assessments, thorough consideration should be given, not only to the current ‘ecological services’ provided by barriers, but also how these may alter under projected climatic scenarios. 33

2) Spatial distribution of barriers and key habitats

Many river systems are highly regulated and contain multiple instream barriers and therefore a thorough spatial assessment of barriers and key habitat types (e.g. refuges and spawning habitats) within watersheds is vital in assessing whether decommissioning will deliver a tangible benefit to the fish fauna (sensu Kemp & O’Hanley 2010). The effects of climate change should be considered in such spatial assessments. For example, habitats that are currently important (e.g. natural refuges) may become degraded or lost through desiccation under future climatic conditions, thus increasing the ecological value of impoundments as artificial refuges. Similarly, the impact of barriers in restricting access to spawning grounds or other key habitat types may reduce in the future if climate change causes the loss or degradation of those habitats.

3) The presence of existing fish passage infrastructure

Barriers that have been previously fitted with fish passage infrastructure are highly unlikely to be considered for decommissioning due to the investment of capital that has already occurred to mitigate barrier impacts. However, climate change could potentially cause hydrological shifts that impact on fishway functionality, or physico-chemical shifts that hamper the ability of migratory fish to utilise fishways, thus rendering them unfit for purpose. The removal of such barriers may become increasingly justified despite past investments having been made in fish passage infrastructure.

4) Characteristics and resistance to climate change of species

The presence of migratory, commercially-important, rare, or protected species strengthens the case for the decommissioning of barriers, particularly if they are incapable of using alternative passage facilities such as fishways. However, if the 34 localised impacts of climate change are likely to result in declines or extirpations of such species, regardless of the impacts of instream barriers, investment in other adaptation measures to mitigate their decline should be prioritised.

2.5 Conclusion

For fishes in drying regions such as south-western Australia, climate change may exacerbate existing stressors, including instream barriers, by increasing the level of fragmentation in rivers, disrupting migratory pathways, reducing the efficacy of existing fishways, and reducing the quality and quantity of both spawning and refuge habitats.

Instream barriers are widely recognised as having deleterious impacts on aquatic ecosystems, and particularly on fish communities; however, in some circumstances barriers may also provide valuable ‘ecological services’. How climate change may influence these contrasting impacts has previously received little attention, but this review has highlighted that this influence is highly likely to be significant, and it will need to be increasingly considered by managers seeking to mitigate the impacts of instream barriers on fishes. The assessment of barrier decommissioning proposals should be underpinned by sound ecological and socioeconomic information and the influence of climate change should be a key consideration within decision making frameworks, thus enabling the impacts, both positive and negative, of barrier decommissioning on fish communities to be properly assessed under future climatic scenarios.

35

Chapter 3

A novel method for identifying fish migration barriers and prioritising

their removal in drying climatic regions

3.1 Introduction

Artificial instream barriers (e.g. dams, weirs, road crossings, culverts) are common in many south-western Australian rivers and provide multiple socioeconomic benefits (e.g. water supply, stream flow gauging, and provision of vehicle access); however, they can also have numerous detrimental impacts on aquatic ecosystems (reviewed in detail in

Chapter 2). Interest in mitigating the impacts of barriers on migratory fishes, via modification, removal or by fishway construction, has gained momentum worldwide in recent years (see below). Like many environmental issues, the scale of this problem vastly outweighs the resources currently being made available to confront it, therefore the development of processes to prioritise instream barriers for remediation works is an area of increasing interest to environmental managers (Kemp & O’Hanley 2010).

To date, numerous processes have been developed, most of which assign a score based on various ecological, physical and financial criteria for each barrier under consideration that is then used to determine a rank order of priority for remediation works under given budgetary constraints (e.g. Pethebridge et al. 1998; Taylor & Love 2003;

WDFW 2009). The main advantage of such score and rank methods lies in their simplicity and speed of application (Kemp & O’Hanley 2010). However, some researchers contend that a lack of coordination between jurisdictional authorities and an over-reliance on these methods have, in the past, resulted in wasted funding through lack 36 of efficiency, erroneous planning and execution of barrier remediation works, and duplication of effort (O’Hanley & Tomberlin 2005; Kemp & O’Hanley 2010).

Kemp and O’Hanley (2010) advocated for the use of more robust optimisation based models, arguing that this holistic and cost-effective approach more efficiently accounted for cumulative effects of multiple barrier networks on habitat connectivity and fish passage within catchments (rather than considering each barrier independently). At a minimum, optimisation modelling requires data for length and/or habitat quality of river reaches between barriers, location and fish passability of barriers, and removal costs

(Kemp & O’Hanley 2010; O’Hanley 2011), but can be adapted to incorporate any number of additional criteria and weighting parameters depending on catchment-specific scenarios and restoration objectives (e.g. Kuby et al. 2005; Zheng et al. 2009). However, the theory and programming requirements underpinning optimisation modelling are highly complex and the lack of ‘turn-key’ applications that are accessible, usable and interpretable to environmental managers has thus far limited their use in real-world applications (Kemp & O’Hanley 2011).

Webb & Padgham (2009) contested that the use of optimisation modelling was not generally warranted, highlighting that solutions apply only to the specific study in question and are generally not widely applicable. Their analysis revealed that, as the overall effects of multiple remediation projects will usually equal their sum effects for fish populations, there is almost never a need to examine interactive effects on multiple modifications (either to stream reach quality or connectivity) and that accurate scoring of individual modifications should be used to prioritise instream barrier mitigation works.

37

3.1.1 Prioritisation methods in practice

A global review of barrier prioritisation processes identified that most had been developed and applied in the United States, Europe, and Australia (Kemp & O’Hanley

2011). In the western United States, a number of protocols have been developed by various agencies for the prioritisation of barrier mitigation projects, mostly aimed at rehabilitating threatened or commercially/recreationally important salmonid populations

(Kemp & O’Hanley 2010). For example, Washington State Department of Fish and

Wildlife developed a comprehensive framework for assessing fish passability and habitat quality around instream barriers that was applied across the State of Washington in 2007 resulting in the identification of ~6,500 culvert barriers, of which, 225 had been removed or modified by 2009 (WDFW 2009). Similar protocols have also been developed in other states (e.g. CSCC 2004; PAD 2009; ODFW 2010). In the eastern United States, barrier inventories have been compiled in an uncoordinated and unreliable fashion with substantial overlap of effort (Kemp & O’Hanley 2010). The US Fish and Wildlife Service also developed a nationwide barrier inventory (GeoFin 2012); however, it was labelled

“simplistic” by Kemp and O’Hanley (2010) as many important ecological and socio- economic factors were omitted.

In Europe, there has been a tendency for stand-alone assessments, generally of large dams, but the focus has widened to include other instream barrier types since the enactment of the EU Water Framework Directive (Kemp & O’Hanley 2010). Barrier inventories have been compiled in Belgium (Monden et al. 2000), the Netherlands (Kroes et al. 2006), Germany (Dumont 2005), and Austria (Zitek et al. 2008), with the information being used to prioritise barriers for remediation works. In England and Wales a standardised barrier assessment methodology is in its infancy but the protocol was also criticised by Kemp & O’Hanley (2010) for, inter alia, relying too heavily on subjective 38 expert opinion rather than objective data. The Scottish Environmental Protection Agency has also recently begun compiling a national barrier database and developed a broad- scale barrier assessment protocol (Kemp et al. 2008).

In Australia, the earliest reported barrier prioritisation process was undertaken in

NSW (Pethebridge et al. 1998). The data collected during this process included physical attributes (e.g. type of barrier, height, width, head loss), weir pool characteristics, stream/catchment characteristics (e.g. distance to proximal barriers, flow regimes, drown- out frequency, habitat condition upstream), fish assemblage information (if known), and a range of other data (e.g. age of structure, owner, status of use, presence of existing fishway). A score for each barrier was calculated based on 11 criteria, which was used to determine the priority ranking for remediation works. This method was later modified to incorporate a more comprehensive dataset during the NSW Detailed Weir Review Project

(NSW DPI 2006).

While officially there has been no centralised, coordinated approach to barrier prioritisation in Australia at a national level, agencies in most Australian states have developed their own assessment and prioritisation methods, most using similar score and rank methods to those of Pethebridge et al. (1998). For example, barrier inventories and prioritisation projects have taken place in Victoria (McGuckin & Bennett 1999; DSE

2003; Ryan et al. 2010), Queensland (e.g. Stewart & Marsden 2006; Carter et al. 2007;

Stockwell et al. 2008; Lawrence et al. 2010; Lawson et al. 2010), and Tasmania (Nelson

2003). In the latter State, no detail was reported on the methods used to determine the prioritisation ranking for the 10 barriers that were ultimately remediated, and despite recommendations for further prioritisation of barriers for removal (Nelson 2003), no further work has taken place, nor is planned there (R. Freeman, Inland Fisheries Service, 39 pers. comm.). In Queensland, barrier remediation has resulted almost exclusively in fishway installations, rather than barrier removals.

The Department of Water (Government of Western Australia) has recently compiled a GIS database of instream barriers and developed a prioritisation protocol for enhancing fish passage that also uses similar assessment criteria to the NSW model, but with a greater emphasis on fish data (e.g. historical vs current distribution, life history information, potential for spread of exotic species through barrier removal) (Storer &

Norton 2011). The protocol consists of a preliminary desktop assessment using the GIS database and existing literature to identify potential fish barriers in an area or catchment of interest, followed by field validation, a cost-benefit analysis, and finally, the prioritisation of candidate barriers; however, the protocol is yet to be applied.

Whilst numerous barrier prioritisation processes have been developed and implemented in several countries and regions, no international or even national coordination or standardisation of these approaches has been undertaken; and importantly, climate change was not specifically considered in any of the prioritisation processes that were reviewed.

3.1.2 Considering synergistic impacts of climate change and instream barriers in prioritisation processes: south-western Australia as a model region

Most native freshwater fishes in south-western Australia are either potamodromous or diadromous (Potter & Hilliard 1986; Pen & Potter 1990; 1991a, b; Morgan 2003;

Chapman et al. 2006; Beatty et al. 2010, 2014a; see also Chapters 1, 4, 5), thus instream barriers represent a key threat to this fauna as they can restrict fish migration both 40 spatially and temporally. This has the potential to limit the reproductive capacity and recruitment levels of populations (Bunn & Arthington 2002); particularly in those species that show a strong correlation between the numbers of spawning migrants and discharge during the annual peak flow period (e.g. Galaxias occidentalis, Nannatherina balstoni,

Nannoperca vittata, Bostockia porosa; Beatty et al. 2014a). Moreover, these negative impacts are likely to be exacerbated by synergistic effects of other threats such as species introductions and anthropogenic climate change (see section 6.5). Future rainfall and surface flow reductions (Hope et al. 2015) threaten to reduce fish passage as instream barriers will drown-out less frequently, further restricting access to key habitat types and exposing fishes that tend to congregate in large numbers below impassable barriers to increased levels of predation (Gregory et al. 2002). Therefore, the region represents an ideal model for examining the interactions between the impacts of instream barriers and climatic drying on freshwater fishes.

Instream barriers can clearly have many negative impacts on migratory fishes, but paradoxically, in some situations they can also provide ecological benefits (see Chapter

2). Ecological ‘services’ provided by barriers, such as halting the spread of an invasive species, can be a key consideration in barrier removal decisions (e.g. Browning et al.

2007). One such potential ecological benefit of instream barriers that has not previously been considered in any documented prioritisation process is the provision of refuge habitat (e.g. impoundments, weir pools) in intermittent river systems. In a drying climatic region such as south-western Australia where natural refuge habitats are expected to decline in both quantity and quality in the future, these artificial habitats are likely to have increasing ecological value as baseflow refuges (see Chapter 2).

The aim of this study, therefore, was to develop a prioritisation process for identifying and assessing instream barriers for remediation or removal that specifically 41 incorporates the potential positive impacts of barriers in a drying climatic region, as well as a range of other ecological and socioeconomic impacts, both positive and negative, which were identified in the review undertaken in Chapter 2. As detailed below, the process consists of five phases (Figure 3.1), and is designed to be applicable at multiple spatial scales in regions undergoing climatic drying. In addition, the study also aimed to trial the process in a pilot study across a range of catchments in south-western Australia

(see section 3.3).

3.2 A barrier identification and prioritisation process for drying climatic regions

3.2.1 Project planning and development

Stakeholder engagement can enhance the quality and success of natural resource management decisions (Reed 2008). Catchment Management Authorities (CMAs) and local environmental organisations are key stakeholder groups that should be included on a steering committee when undertaking instream barrier mitigation/removal projects as they usually have strong local/regional community links. The committee should also include representatives from appropriate governmental departments, landholder/irrigation groups, traditional owner groups, as well as experts in environmental engineering, aquatic ecology and a suitably experienced project manager to guide the project from conception through to completion.

The spatial scale of the project is designated during this initial phase of the process with the choice of region(s) or, if coverage is to be more spatially constrained, catchment(s) in which to conduct the process largely dependent on the resources

Barrier identification and prioritisation process for drying climatic regions 42

Figure 3.1. Flow chart of the process developed for identifying, assessing and prioritising instream barriers for remediation/removal

in drying climatic regions. N.B. steps in bold typeface were completed during pilot studies as detailed in this chapter. 43 available (human and monetary), as well as the conservation issues and objectives relevant to the target area. For instance, catchments that house commercially-important and/or threatened taxa, particularly obligate migratory species, are likely to be of greater interest; however, other political, socioeconomic, and ecological factors may also influence the project’s scale and objectives, as will the jurisdictional boundaries of the relevant waterway management authorities involved in the project.

Information on the hydrology, ecology, and instream barriers of the target area, including any projected or likely impacts of climate change, should then be collated during a desktop review of available literature and data. A thorough review of data on the geospatial distribution and characteristics of instream barriers in the target area, often maintained by governmental water management agencies, is a vital part of the process.

Spatial and ecological information on the aquatic fauna within the target area, particularly seasonal patterns of migration, is crucial and should be matched against the hydrology of the system in order to guide several aspects of the score and rank system (see below).

Additionally, migration patterns can sometimes vary among sympatric species in space and time (Figure 3.2) and this should be considered during the planning and execution of the process.

Landholders, particularly those with riparian frontage in the catchment(s) of interest, are important stakeholders in this process and should be contacted early in the project development phase. This can be done in a variety of ways (e.g. mail, e-mail, telephone, community workshops, etc.) in order to provide information about the process, negotiate access to barriers of interest, and to give landholders an opportunity to contribute additional data to the project (see Appendix 1). An assessment must then 44

Figure 3.2. Seasonal migratory movements (top) of fishes of the Goodga River, south-

western Australia: green = adults, blue = larvae, red = juveniles. Arrows

indicate upstream and downstream movements. Also presented (bottom) is

the mean monthly historical discharge within the system (discharge data

from Department of Water, Government of Western Australia).

45 be made of key knowledge gaps that, if not addressed, will hinder the completion of the next two phases of the project. These are likely to be related to: 1) a lack of information on the location of barriers and/or refuge habitats; and, 2) a lack of information on the aquatic fauna within the study area. Knowledge gaps, as determined by relevant experts, either on or engaged by the steering committee, should be addressed before progressing to the next phase of the project.

3.2.2 Validation of barrier information

Once adequate information has been gathered to proceed with the process, aerial and/or ground-truthing of barriers and refuge habitat data should be undertaken. In some areas

(e.g. predominantly deforested catchments), aerial imagery may be used to validate the location of instream barriers; however, this method is likely to be ineffective when assessing small streams and rivers with intact riparian vegetation cover. Low-altitude (i.e.

<150 m above ground level) aerial survey is advocated by Fausch et al. (2002) as an effective method for rapidly assessing stream habitat features, and is especially useful for validating barrier information, and also recording and mapping features (e.g. instream barriers (artificial or natural) and refuge habitats) that were not identified during the desktop review phase. Either of the two truthing methods (i.e. aerial or ground) may be more economically viable or suitable depending on a range of factors, including the size and complexity of the target area, the extent of road access, and the proximity of the target area to the headquarters of aerial survey operators. A combination of both methods is likely to prove most effective (e.g. rapid aerial assessment followed by more detailed and targeted ground-truthing). 46

Physical and hydrological characteristics (e.g. structural dimensions, head loss, whether the barrier is run-of-the-river or undershot by pipework, presence/absence of bypass channels) of candidate barriers identified during the desktop review, as well as those of any additional barriers encountered, can usually be assessed during aerial validation surveys. Any barriers that feature a head loss that does not exceed the leaping ability of native fishes can be identified during this initial screening and excluded from further assessment. Note that water levels can introduce uncertainty in observations; therefore, the timing of surveys should coincide with periods when fishes are migrating within the target area (i.e. typically during winter/spring flows in south-western

Australia). Any observations obscured by vegetation cover or adverse light conditions can be noted and these barriers prioritised for follow-up ground-truthing if necessary.

Secondary aerial surveys during the annual baseflow period should also be undertaken in intermittent river systems, as this allows the spatial distribution of permanent refuge habitats (often a major knowledge gap) to be mapped.

3.2.3 Barrier prioritisation

Instream barriers identified during the previous phase are then ranked in order of priority for remediation works or removal in order to enhance fish passage. Each barrier is assessed using a comprehensive score and rank matrix (Table 3.1). Criteria are weighted according to their relative impact on fish migrations within target catchments. For example, the proportion of the flow period during which the barrier is drowned-out (i.e. passable) is weighted highly as it has a much larger impact on fish migration than lower

Table 3.1. Barrier prioritisation score calculation matrix (adapted from Pethebridge et al. 1998; Ryan et al. 2010; Norton & Storer 2011).

POSITIVE CRITERIA Weighting Rating Score 3 2 1 0 -1 (Weighting x Rating) 1: Fish present (or expected) within stream section Threatened fish 5 migratory migration likely local movements not present n/a 5 x rating only Native fish diversity (% of species in catchment 5 >80% 50-79% 25-49% <25% n/a 5 x rating represented)

2: Barrier location and relative size Increase in upstream access to next barrier if 5 >15% 9-15% 2-8% <2% n/a 5 x rating barrier mitigated (% of total catchment stream length) Percentage of total stream length upstream of 3 >80% 50-79% 25-49% 5-24% <5% 3 x rating barrier (regardless of barriers) Percentage of total fish barriers in catchment 3 <10% 10-49% 50-79% >80% n/a 3 x rating located downstream Fish access through drown-out (% flow period) 10 <20% 20-34% 34-49% 50-75% >75% 10 x rating

3: Habitat quality and availability upstream of barrier Quality of riparian buffer upstream of barrier 2 Excellent Good (mostly Moderate Poor (mostly n/a 2 x rating (near pristine) undisturbed) (patchy/semi- degraded) degraded) Water quality upstream of barrier (i.e. within the 2 High Slightly Moderately Highly exceeding 2 x rating ANZECC (2000) guideline trigger values for conservation disturbed disturbed disturbed trigger value various protection levels) value threshold Increase in upstream refuge access if barrier 8 >30% 15-29% 5-14% <5% n/a 8 x rating mitigated (% of total refuge habitat in catchment) Percentage of year flowing at barrier 2 100% 40-99% 5-39% <5% n/a 2 x rating

Positive criteria SUB-TOTAL (max. 135) 47 Table 3.1. (cont.) Instream barrier prioritisation score calculation matrix.

48 NEGATIVE CRITERIA Weighting Rating Score 3 2 1 0 (Weighting x Rating) 4: Negative impacts of barrier mitigation Barrier preventing spread of Alien fish -10 confirmed likely possible unlikely - 10 x rating Lack of stakeholder support -1 confirmed likely possible unlikely - 1 x rating

Negative criteria SUB-TOTAL (max. -33) GRAND TOTAL = Positive criteria score + Negative criteria score (max. 135)

SUPPLEMENTARY SCORE CRITERIA*

* Additional negative impacts subtracted from Grand Total under a barrier removal scenario (*apply these scores only when a barrier pool provides a refuge habitat in baseflow) Biodiversity value of barrier pool as a baseflow -3 (threatened spp. (50-79% of all (25-49% of all (<25% of all fish - 3 x rating refuge habitat present and/or fish spp.) fish spp.) spp.) >80% of all fish spp.) Species uniqueness in barrier pool (compared to -10 2 or more spp. 2 or more spp. 1 sp. (not zero uniqueness -10 x rating other known refuges) including at least (none threatened) 1 threatened threatened) Relative proportion of total baseflow refuge habitat -12 >15% 10-14% 5-9% <5% - 5 x rating created by barrier Proximity to nearest refuge -2 >5 km 2-5 km 1-2 km <1 km - 2 x rating Major downstream sedimentation risk -4 confirmed likely possible Unlikely - 4 x rating Supplementary score criteria SUB-TOTAL (max. -93)

SUPPLEMENTARY GRAND TOTAL = GRAND TOTAL + Supplementary score criteria SUB-TOTAL (max. 135) 49 weighted criteria, such as the quality of habitat upstream of the barrier (see Table 3.1).

Access to refuge habitat is another highly weighted criterion, reflecting one of the key aims of the process in delivering tangible ecological benefits through the remediation/removal of instream barriers.

Criteria pertaining to hydrology (e.g. barrier drown out frequency, percentage of year flowing at barrier, aquatic refuge availability upstream of barrier) and ecology (e.g. quality of riparian buffer and water quality upstream of barrier, conservation status of fish fauna) can be scored and assessed under various altered state scenarios where robust projected data are available. In drying climatic regions, the availability of refuge habitat is likely to diminish in the future, and artificial aquatic refuges created by instream barriers (e.g. reservoirs, weir pools) may become increasingly important for aquatic communities. One of the key novel aspects of this prioritisation process is the inclusion of supplementary scoring criteria for barriers that have an associated impoundment that provides baseflow refuge habitat. In these instances, the relative ecological value of the artificial refuge cf. other refuges in the target catchment is assessed (Table 3.1). The negative weighting of the supplementary scoring criteria has the effect of downgrading the score and prioritisation rank of barriers with associated refuge habitats of high ecological value (Table 3.1). Such barriers are not likely to be considered for removal, as this would potentially lead to the loss of the refuge habitat, but they may be considered for remediation by alternative means (e.g. fishway installation) should the prioritisation ranking warrant such action.

Following ranking, the prioritised list of barriers is presented to stakeholders for feedback. This crucial phase of the process must necessarily include direct consultation with barrier owners/managers to workshop prioritisation strategies. Any socio-economic factors (e.g. goods and services, local amenity, recreational opportunities, heritage values

50 provided by barriers) surrounding the barriers that have been prioritised can be discussed at this stage of the process in order to refine the final selection of barriers for remediation/removal works. Any candidate barriers that are identified as redundant may be shortlisted for removal, whereas, high-priority non-redundant structures are likely to only be considered for remediation (usually via fishway construction) in order to maintain their utility. The number of barriers selected for removal or remedial works will, of course, depend on budget availability and project objectives.

3.2.4 On-ground works and monitoring

Following agreement between the steering committee and stakeholder groups over proposed barrier removal/remediation plans and the securing of necessary work permits, the on-ground works phase of the process can begin. However, ecological monitoring should be undertaken to collect appropriate baseline data prior to the commencement of any site works (Stewart & Marsden 2006). It is crucial that appropriate resources are made available for ecological monitoring in order to determine the success of the project.

The scope and scale of ecological monitoring will be dependent on the barrier, the aquatic ecosystem, and the level of existing information on that ecosystem. Sediment quality, including particle size distribution, and load in the reservoir (should one exist) upstream of the barrier should also be assessed to determine the potential for downstream sedimentation of habitat or the release of any toxic or nutrient-enriched sediments.

Monitoring should be undertaken over an appropriate timeframe so that any ecological impacts of the barrier remediation/removal (e.g. extension of distributional range, enhancement of population connectivity, increases in levels of recruitment and relative abundance of migratory species) can be objectively documented and quantified. 51

Broadly speaking, the downstream and upstream aquatic fauna community, water quality, and sediment loads should be assessed using established survey techniques such as those used in the development and monitoring of biotic indices of ecosystem health (e.g.

Chessman 2003). Monitoring should be undertaken on a seasonal basis for a period of at least one year before, and at least two years after the completion of on-ground works.

Barrier removal/remediation works should be undertaken by suitably qualified contractors under the supervision of structural and/or environmental engineers to ensure safety and efficiency. Appropriate consideration should also be given to minimise disturbance of riparian zones, instream habitats and the surrounding environment more generally (e.g. introduction of disease, alien species or pathogens). Depending on the scale of the project and the location of barriers, some degree of site rehabilitation may be required. Where the re-establishment of vegetation following instream barrier removal is left to occur without human intervention, newly exposed riparian areas can be susceptible to colonisation by invasive plant species (e.g. Shaforth et al. 2002; Orr & Stanley 2006).

Therefore, more proactive restoration protocols such as direct seeding or planting with a diverse range of locally native plant species (especially fast growing, early successional species) is preferable in order to minimise the risk of invasive species (Wolden &

Stromberg 1997). The nature and scale of the rehabilitation should be decided by the steering committee with input from CMA representatives and environmental engineers.

This may include both physical and ecological rehabilitation of the streamline.

Minimising the impact of sediment transport during removal projects can be achieved by appropriate timing and pacing of removal, along with other mechanical stabilisation approaches such as grading or armouring river banks with geotextiles or riprap (Randle

2003; Rathburn & Wohl 2003).

52

3.2.5 Project extension

As with all natural resource management projects, every effort should be made to ensure that the project is promoted widely to maximise the public education benefit, as increasing awareness of the factors impacting natural systems is crucial in their ongoing conservation (Fien et al. 2001). The steering committee should formulate a project extension plan that includes stakeholder workshops and information sessions (as previously mentioned), as well as extension to a wider audience through media releases and other appropriate channels.

Demonstration sites for restoration projects are widely used to raise awareness and community support (e.g. Stewart & Marsden 2006) and it is recommended that barrier removal/remediation sites be considered for use as demonstration sites if deemed appropriate. This may involve erecting signage with information on ecological conditions at the site before and after on-ground works, as well as inviting local stakeholders, media, and interested members of the local community to attend and observe removal works, should this be feasible.

3.3. Pilot study

A full trial of the process for identifying, prioritising and remediating instream barriers as outlined in this chapter could not be undertaken due to the lack of a mandate and funding for such work. Instead, a pilot study was conducted in order to apply and test some of the components of this process. The pilot study took place in six catchments in south-western

Australia that were specifically selected to represent varying degrees of anthropogenic disturbance (most notably through clearing of native vegetation for agriculture). All of 53 the selected catchments had considerable data available on the spatial distribution of fishes, and most housed populations of rare or listed threatened species (e.g. Galaxias truttaceus, Galaxiella munda, Nannatherina balstoni, Nannoperca pygmaea).

3.3.1 General methods

A review of geospatial data on potential instream barriers to fish migration was initially conducted in the selected catchments using the instream barrier database maintained by the Department of Water (DoW), Government of Western Australia. This database includes information on infrastructure that disrupts longitudinal stream connectivity completely (e.g. dams, weirs) or partially (e.g. culverts, bridges, fords, floodgates, levees, etc.) (Norton & Storer 2011). The DoW uses an automated process to assign a level of impact on fish movement for each barrier according to the barrier type and various characteristics of the river in which it is located (e.g. major vs minor, perennial vs ephemeral). This process produces an impact code that ranges from 1 (nil priority) to 6

(very high priority) which can be useful for highlighting significant barriers to fish migration (Norton & Storer 2011).

A number of barriers identified in the desktop review were pre-selected for further investigation i.e. structures with an impact code > 3 that were located on the main river channel. Barriers rated with an impact code ≤ 3 comprise bridges, culverts, and other structures that are unlikely to pose major impediments to fish migration (see Tables 2 and

4 in Norton & Storer 2011), hence, they were excluded from any further assessment. Note however, that some lower impact code barrier types were checked during initial aerial surveys to validate this assumption. Additionally, information flyers and questionnaires were mailed to riparian land holders (whose contact details had been attained from

54

CMAs) in three of the six catchments. This was undertaken in order to raise awareness of the issue of instream barriers and gather basic supplementary data on barriers and refuge habitats (see Appendix 1).

Validation of data is an important component of the prioritisation process as the

DoW database contains information that has not been field validated and lacks data for natural barriers such as waterfalls and cascades (Norton & Storer 2011). As road access to rivers was limited in the target catchments, aerial survey by chartered helicopter was used to validate barrier data.

3.3.2 Aerial surveys

Aerial surveys took place in October 2012 (i.e. mid-spring) in the Ludlow, Mitchell and

Goodga rivers, and December 2014 in the Fly Brook, Scott, and Margaret catchments

(Figure 3.3). Surveys coincided with ‘shoulder’ flow periods to avoid difficulties associated with assessing barriers that may be drowned out during ‘peak’ flow conditions.

Surveys were conducted from a Robinson R44 helicopter (Figure 3.4) that was flown

125-150 m above ground level (AGL) at an average speed of ~40 km.h-1 over the target catchments. A GPS device (Garmin eTrex Legend HCx) electronically tethered to a laptop computer running the mapping software OziExplorer was used for inflight navigation. This equipment was operated by the author who conveyed verbal navigational instructions to the pilot over the helicopter’s internal communications system. The flight path was logged and overlayed in real‐time on a digital 1:50 000 topographic map using

OziExplorer software. All preselected barriers were saved as waypoints in OziExplorer to aid inflight navigation.

55

Figure 3.3. Locations of the six target catchments in which the pilot study was

undertaken to trial the barrier identification and prioritisation process in

south-western Australia.

Figure 3.4. Helicopter used to conduct aerial surveys in the pilot study (left), and

aerial view of Ten Mile Brook Reservoir in the catchment,

south-western Australia (right).

56

A handheld voice recorder (Olympus VN-712PC) was used to record a range of data for all preselected barriers including barrier type and function, construction material, head loss, and the nature of any flow at the site. Stream width, accessibility of the barrier site, surrounding land use(s), and whether the barrier appeared to be redundant were also recorded (see Appendix 2). Data were also recorded for any significant fish barriers that were encountered during the survey that had not been preselected, including any significant natural barriers (e.g. waterfalls, cascades). Digital images were also taken of all barriers that were encountered.

Aerial surveys were also conducted in two of the target catchments (Mitchell,

Goodga) to search for baseflow refuge habitats. These surveys used the same methods described above, but were conducted during a baseflow period (March 2013). This was undertaken as a trial to address a key knowledge gap for the majority of rivers in south- western Australia, with the data gathered contributing directly to the calculation of barrier prioritisation scores in these catchments.

3.3.3 Prioritisation process

Data from the aerial surveys were compiled in an Excel spreadsheet and a prioritisation score was calculated for each barrier using a scoring matrix (Table 3.1). It should be noted that due to insufficient funding, data for some of the lower weighted criteria (i.e. water quality, stakeholder support for barrier removal, and sedimentation) could not be gathered and were thus excluded from the calculations. Also, in some catchments there were spatial gaps in the dataset used to calculate scores for criteria pertaining to fish distribution (Scott, Ludlow, Fly) and baseflow refuge habitat (Ludlow) for some barriers. Normally, such knowledge gaps would be addressed before proceeding 57 with the calculation of prioritsation scores, but for the purposes of this pilot study, scores were inferred from data available elsewhere in the relevant catchment for fish distribution criteria, and on available aerial imagery and stream flow data for baseflow refuge criteria.

It should also be noted that one instream barrier in the Goodga River catchment (barrier

#49) that has a functional vertical-slot fishway (see Chapter 4 for further information) was scored as if it were a non-remediated barrier in order to ascertain its prioritisation ranking relative to other barriers assessed in the pilot study.

3.4 Results

In total, 57 high priority barriers (i.e. impact code > 3) were identified from the DoW barrier database in the six target catchments; however, 50.9% of these preselected features did not have a significant head loss (i.e. ≤ 0.3 m) and were therefore assessed not to be significant fish barriers (Table 3.2). Aerial validation surveys revealed 36 additional anthropogenic fish barriers (and eight natural barriers) across the six catchments (Table

3.2). Some of these structures were registered in the DoW database as lower priority barriers (i.e. impact code ≤ 3) and therefore had not been preselected for aerial survey

(see rationale in Methods); however as expected, the vast majority of lower impact code barriers had insufficient head loss to be considered significant barriers to fish migration.

The number of significant fish barriers per surveyed stream kilometre across all catchments was 0.27 (Table 3.2), i.e. about one barrier every 4 km. Fly Brook and the

Goodga River had more artificial fish barriers per surveyed stream kilometre than the other catchments (Table 3.2). This reflected the higher intensity of agricultural land use in these catchments (and consequently greater frequency of onstream dams and fords), although note that lower order tributaries (which tend to have a higher prevalence

58

Table 3.2. Summary of key results of the desktop review and aerial survey phase of the barrier identification process conducted within the six

target catchments. NB - natural barriers are not included in this summary.

Catchment Number of Number of Number of Number of Length of stream Number of preselected high preselected barriers additional potentially aerially surveyed artificial priority fish that were not barriers significant fish (km) barriers per barriers (impact significant fish identified during barriers aerially surveyed stream code > 3) barriers aerial survey surveyed km A B C = (A – B) + C

Margaret R. 11 9 4 6 48.5 0.12

Scott R. 3 1 3 5 24.1 0.21

Fly Bk 14 3 3 14 23.5 0.60

Ludlow R. 19 12 9 16 72.7 0.22

Mitchell R. 9 4 2 7 41.3 0.17

Goodga R. 1 0 15 16 23.7 0.68

TOTAL 57 29 36 64 233.8 0.27

A) B)

Figure 3.5. Instream barriers to fish migration assessed by aerial survey in the A) Margaret River, B) Scott River, C) Fly Brook, D)

Ludlow River, E) Mitchell River, and, F) Goodga River. Fish distributions are presented as pie charts and the proportion of

baseflow refuge habitat in various sections of the catchments (separated by dashed lines) is indicated in blue boxes.

59

60

C) D)

Figure 3.5 (cont.). Instream barriers to fish migration assessed by aerial survey in the A) Margaret River, B) Scott River, C) Fly Brook, D)

Ludlow River, E) Mitchell River, and, F) Goodga River. Fish distributions are presented as pie charts and the proportion of

baseflow refuge habitat in various sections of the catchments (separated by dashed lines) is indicated in blue boxes.

E) F)

Figure 3.5 (cont.). Instream barriers to fish migration assessed by aerial survey in the A) Margaret River, B) Scott River, C) Fly Brook, D)

Ludlow River, E) Mitchell River, and, F) Goodga River. Fish distributions are presented as pie charts and the proportion of

baseflow refuge habitat in various sections of the catchments (separated by dashed lines) is indicated in blue boxes. 61

62 of instream dams in south-western Australia cf. main channels) were not aerially surveyed in some of the catchments (e.g. Margaret, Scott).

In total, 48.5 km of the Margaret River mainstem was aerially surveyed in the middle and upper reaches of the system (Figure 3.5A). Two significant fish barriers were identified during the study; a weir with moderate head loss (barrier #1, Rendall Close

Weir) and a road crossing with a single pipe culvert (barrier #5). Four artificial structures that did not appear in the DoW database were identified in this section (barrier #2, #3, #4,

#6), but all structures were assessed as unlikely to have a significant impact on fish migration due to insufficient head loss (Figure 3.5A).

In the Scott River, a total of 24.1 km of streamline was surveyed for instream barriers upstream of the bridge on Scott River Road, with five potential fish barriers identified (Figure 3.5B). Three of these were low head loss fords in the upper part of the catchment (Figure 3.5B) that were assessed as unlikely to represent a significant barrier to fishes due to the high frequency of inundation during the annual flow period. Two barriers in the Scott River were assessed as likely to have a significant impact on native fish migration (Figure 3.5B) including the v-notch flow gauging weir at Brennan’s Ford

(barrier #8). The head loss of this barrier was ~0.5 m and it was located in a section of the catchment with a substantial amount of refuge habitat both downstream and upstream.

Fourteen potential fish barriers were identified in Fly Brook, most of which were onstream farm dams in the headwaters of the catchment (Figure 3.5C). Whilst these structures were identified as significant fish barriers, many had low-gradient spillway channels that would permit fish passage during high flow periods. Two barriers (#13 and

#14, Figure 3.5C) were road crossings featuring large diameter pipe culverts that would 63 be passable to fish during most months of the year in this perennially flowing, groundwater-fed system. The only significant barrier in the lower part of Fly Brook was the v-notch gauging weir (barrier #12), which had a head-loss of ~0.4 m and was located approximately 2 km upstream of the confluence with the Donnelly River (Figure 3.5C).

The largest catchment surveyed during the pilot study was the Ludlow River. In total, 72.7 km of streamline was aerially surveyed, including a number of major tributaries, and 16 potential barriers were identified (Figure 3.5D). The majority of significant fish barriers in the Ludlow system were clustered on a single tributary (Tiger

Gully) flowing through a sand mining area in the lower reaches of the catchment (Figure

3.5D). Most of the remaining barriers identified, particularly those in the upper reaches, featured low head loss and were not likely to be major impediments to fish migration.

The Mitchell River catchment is contained entirely within a conservation estate

(i.e. Mt. Lindesay National Park) and had no major instream barriers in its lower reaches.

However, four significant instream barriers were identified in the middle and upper reaches, including a v-notch gauging weir with a head loss of ~0.5 m (Figure 3.5E).

Numerous natural barriers (waterfalls and cascades) were also observed in this catchment

(Figure 3.5E). The Goodga River, in contrast, has a more extensively modified catchment featuring 16 potentially significant fish barriers (Figure 3.5F), including four dams

(barrier #50, #51, #62, #63), and a number of pipe culverts with moderate head loss.

3.4.1 Barrier prioritisation

Of the six surveyed catchments, Fly Brook had the lowest mean barrier prioritisation score of 14.4 ± 4.4 S.E. (i.e. ~11% of the theoretical maximum score) reflecting the

64 abundance of low scoring barriers in its upper reaches. The Ludlow River also had a low mean prioritisation score of 20.4 (± 3.2 S.E.). Mean scores were > 40 in all other catchments, and highest in the relatively pristine Mitchell River (54.7 ± 12.2 S.E.; i.e.

~42% of the theoretical maximum score).

Two of the top three ranked barriers were in the Mitchell River catchment (Table

3.3). The highest scoring barrier was a v-notch gauging weir in the upper catchment

(barrier #44, score = 113; Figure 3.6), and the third highest a ford located just upstream of the confluence with the Hay River (barrier #42, score = 87) (Table 3.3). V-notch gauging weirs in the Goodga River (barrier #49, score = 112, ranked 2nd; Figure 3.6), Scott River

(barrier #8, score = 76, ranked 5th; Figure 3.6), Fly Brook (barrier #12, score = 61, ranked

10th; Figure 3.6) were also among the highest ranking instream barriers (Table 3.3).

Margaret River contained only one of the top ranked barriers: a rock weir in the middle reaches (barrier #1, score = 85, ranked 4th; Figure 3.6) (Table 3.3). The only catchment without a top 10 ranking barrier was the Ludlow; the highest scoring barrier in this system was a concrete drop-structure used for erosion control and stream gauging (barrier #26;

Figure 3.6) with a score of 57 (ranked 14th).

Seven of the highest ranking barriers were also found to provide baseflow refuge habitat and therefore had supplementary scoring criteria applied in the prioritisation process (Table 3.3). The ranking order remained mostly the same for the top-ranked barriers following application of supplementary scoring with one major exception: the score for the top ranked barrier (i.e. #44 in the Mitchell River) was reduced by 49 points, dropping it from 1st to 6th in the priority ranking order (Table 3.3). This substantial change in ranking reflects the high conservation value of the refuge habitat provided by the weir pool associated with this barrier. 65

Table 3.3. Highest priority instream barriers across six target catchments (Ludlow,

Margaret, Scott, Fly Brook, Mitchell, Goodga) for consideration of

remediation/removal to enhance native fish migration. The supplementary

score is given in parenthesis for barriers that provide refuge habitat in their

associated impoundment (see section 3.2.3 for rationale). #NB- the Goodga

gauging weir has a functional vertical-slot fishway, but was assessed as if

this did not exist (see above).

Rank Catchment Barrier name Score (barrier #) (supplementary score)

1 Mitchell Beigpiegup gauging weir (#44) 113 (64)

2 #Goodga Goodga gauging weir (#49) 112 (103)

3 Mitchell Ford near Hay confluence (#42) 87

4 Margaret Rendall Close Weir (#1) 85 (79)

5 Scott Brennan’s Ford gauging weir (#8) 76 (70)

6 Goodga Pipe culvert on farmland (#60) 67

7 Scott Ford in farmland (#9) 64

8 Scott Scott River Road redundant ford (#7) 63 (57)

9 Goodga Onstream dam (#50) 63 (54)

10 Fly Brook Fly Brook gauging weir (#12) 61 (52)

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Barrier #1 Barrier #8

Barrier #44 Barrier #49

Barrier #12 Barrier #26

Figure 3.6. The highest ranking barrier in each of the six target catchments assessed

during the pilot study: Barrier #1 – Rendall Close Weir (Margaret River,

ranked 4th), Barrier #8 – Brennan’s Ford gauging weir (Scott River, ranked

5th); Barrier #44 – Beigpiegup gauging weir (Mitchell River, ranked 1st);

Barrier #49 – Goodga gauging weir/fishway (Goodga River, ranked 2nd);

Barrier #12 – Fly Brook gauging weir (Fly Brook, ranked 10th); Barrier

#26 –drop structure used for erosion control and stream gauging (Ludlow

River, ranked 14th). 67

3.5 Discussion

There is an increasing trend worldwide of instream barriers being removed or remediated in order to enhance habitat connectivity for migratory fishes yet none of the associated processes for prioritising these works have specifically incorporated climate change in barrier assessments (see Chapter 2). In south-western Australia, despite the prevalence of potamodromy among native freshwater fishes, on-ground work in this region has so far been limited to the installation of a just a few fishways (e.g. Morgan & Beatty 2006;

Beatty et al. 2007, 2014b). In a positive first step, the Western Australian Government’s

Department of Water has recently addressed the lack of work in this field through the development of a fish barrier database and assessment/ranking protocol for the State

(Norton & Storer 2011).

The current study has greatly expanded upon this work by developing a score and rank protocol that includes assessment criteria pertaining to the impact of barriers on connectivity of baseflow refuge habitats, as well as more traditional barrier prioritisation criteria (e.g. conservation status of resident fauna, drown-out frequency, increase in upstream access, habitat quality, and potential for alien species dispersal). This process also includes assessment criteria that accounts for the biodiversity value of any refuge habitats associated with instream barriers (e.g. reservoirs, weir pools) that may be threatened under a barrier removal scenario. Furthermore, the process is easily able to incorporate projected hydrological data that can potentially influence passability of instream barriers as well as the temporal availability and permanence of baseflow refuge habitats under a range of future climate scenarios. The spatial distribution and habitat characteristics of baseflow refuges and instream barriers drive the distribution and community structure of fish assemblages in intermittent river systems (Magoulick &

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Kobza 2003; Perry & Bond 2009; Beesley & Prince 2010). In drying climatic regions, the amelioration of the effect that barriers have in disrupting the connectivity of baseflow refuges in intermittent rivers will become increasingly important (sensu Bond et al. 2008) in the conservation of resident fishes, particularly in refuge limited systems.

3.5.1 Aerial survey methods

The application of an instream barrier identification and prioritisation process in the pilot study was the first of its kind to be undertaken in south-western Australia. This study presented an opportunity to test the effectiveness of aerial survey via helicopter for identifying and assessing instream barriers and refuges. Aerial survey is used across many biological disciplines to census faunal groups such as waterbirds (e.g. Kingsford &

Porter 2009), feral livestock (e.g. Walter & Hone 2003), and marine fauna (e.g. Marsh &

Sinclair 1989). Bond (2007) successfully trialled the use of helicopters to map aquatic refuge habitats in south-eastern Australian rivers and this technique proved equally effective for the rivers surveyed during the pilot study. Currently, there are few viable alternatives for determining the spatial distribution of critical baseflow refuges (that are often very small; sometimes only a few metres wide) in intermittent river systems in forested regions of southern Australia. However, ongoing advancements in remote sensing technology (see for example Hilldale & Raff 2008; Marcus & Fonstad 2010;

Klemas & Pieterse 2015) may complement or eventually replace manual aerial surveys for the mapping of small refuge pools.

Although aerial survey costs appear high (i.e. ~$1,000 per hour at time of publication), they compare favourably to the alternative of ground-based truthing due to the time efficiency involved. To illustrate this point, a total flight time of approximately 69

15 hours (undertaken over three days) yielded data for 64 potential fish migration barriers and screened a further 125 barriers that were deemed not to be significant barriers. To survey this number of barriers by vehicle would have required a substantially greater time investment, and would not have achieved an equivalent spatial coverage due to the lack of vehicle access throughout large sections of the target catchments. The discovery of additional barriers (both artificial and natural) that were missing from the DoW database further highlights the value of aerial survey as a tool for validating and expanding upon data collated in the desktop review phase of the process.

Aerial survey was not without its shortcomings, however, with some barriers proving difficult to assess due to obscured visibility through dense riparian vegetation.

The cost-effectiveness of aerial survey also reduces with increasing distance between target catchments and the base of charter flight operations. Aerial surveys should therefore attempt to minimise the time spent transiting between target catchments in order to maximise cost-efficiency. It is also crucial that local riparian landholders are notified of dates and times of planned flights well in advance as a courtesy to allow sufficient time for livestock to be moved away from river channels in order to prevent them from being startled by low-flying aircraft. Despite these shortcomings, aerial survey proved to be a highly effective means of compiling data to address key knowledge gaps in order to progress the barrier prioritisation process.

3.5.2 Stakeholder engagement

Acquisition of data on instream barriers and refuge habitats via questionnaire responses from riparian landholders was trialled in three of the target catchments. Local CMAs

(also an important stakeholder group in the prioritisation process) provided landholder

70 contact information, resulting in the collection of basic data for only three instream barriers and four permanent pools in two catchments. Significantly, less than a third of the landholders that were contacted responded to the questionnaires, demonstrating that this method cannot be solely relied upon to compile spatially comprehensive datasets.

Rather, it can provide complementary data to help address what are, more often than not, crucial knowledge gaps in southern Australian rivers, i.e. the spatial distribution of fish refuge habitats and instream barriers. Contacting landholders in this way is also a method of engaging these important stakeholders in the barrier prioritisation process.

No formal guidelines for decommissioning instream barriers exist in Australia; however, examples are available that demonstrate the importance of involving a broad range of stakeholders in the development of barrier removal projects to ensure success and avoid opposition and conflict (I&I NSW 2009; Lejon et al. 2009). The importance of stakeholder engagement and education in all aspects and at all stages of dam decommissioning cannot be overstated. This engagement should be driven not only by scientists involved in these projects, but also resource agencies, extension services, NGOs and dam owners themselves to ensure that a broad range of stakeholders are engaged and empowered throughout the decision making process (Sarakinos & Johnson 2003). Barrier removals can evoke strident opposition amongst some stakeholders (Lejon et al. 2009), therefore careful multi-disciplinary planning, monitoring, and effective public communication strategies need to be implemented throughout the prioritisation and removal process in order to minimise potentially negative ecological and socio-economic impacts resulting from these decisions.

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3.5.3 Prioritisation of instream barriers

All but one of the target catchments in the pilot study housed listed threatened fishes with potamodromous life cycles that predispose these species to negative impacts of instream barriers during annual breeding migrations (see review in Chapter 2). These catchments were deliberately chosen, as a full “real-world” application of the barrier prioritisation process would most likely focus on catchments housing species that are legislatively recognised as requiring conservation management. Indeed, the sensitivity of these species to numerous anthropogenic stressors was reflected in the relative infrequency of significant fish barriers encountered in these catchments.

The catchments that had a combination of a high diversity of threatened fish species and relatively few significant instream barriers (i.e. Mitchell, Margaret) had the highest mean barrier prioritisation scores. In contrast, catchments with few or no threatened species and a relatively high number of instream barriers (i.e. Ludlow, Fly

Brook) tended to have lower mean prioritisation scores, reflecting their higher degree of modification and anthropogenic disturbance. An exception to this was the Goodga River which featured extensive catchment modification and a high number of significant fish barriers, yet had a relatively high mean prioritisation score, which was explained by the unusually high diversity of threatened fish species (i.e. G. truttaceus, G. munda, N. balstoni) inhabiting this catchment.

Barriers located in the lower reaches of catchments ranked highly in the prioritisation process. This was due to the high species richness within the lower reaches of the chosen target catchments, as well as the potential for proportionately larger habitat gains for resident species when barriers in the lower reaches of catchments are removed or remediated. An exception to this; however, was the top ranked barrier in the study, a v-

72 notch gauging weir (barrier #44, see figures 3.5E, 3.6) that was located in the upper reaches of the Mitchell River. The high score (87.6% of the theoretical maximum score) for this barrier resulted from a combination of attributes, i.e. a high diversity of threatened species, low drown-out frequency, and restriction of access to a relatively high proportion of stream habitat, including, most importantly, almost one third of the total baseflow refuge habitat available in the catchment.

The relative importance of artificial refuge habitats is expected to increase as the climate in south-western Australia becomes drier in the future due to anthropogenic climate change. A total of 25 barriers (including six of the top 10 ranked barriers) were identified in the target catchments as providing refuge habitats in their associated impoundments and, hence, were subjected to supplementary scoring criteria designed to downgrade prioritisation scores where the conservation value of barrier-associated artificial refuges is high. Supplementary scoring is applicable only in a barrier removal scenario and does not apply in cases such as fishway installations where the refuge habitat provided by the barrier remains in place. In the case of the top-ranked Mitchell

River gauging weir, the supplementary score severely downgraded its final prioritisation ranking, indicating that removal of this barrier would be much less desirable than the installation of a fishway.

The high ranking of the v-notch weir on the Goodga River (barrier #49, ranked

2nd) justified the earlier construction, in 2003, of a vertical-slot fishway at this barrier site, which facilitated the range expansion of an endangered population of G. truttaceus

(Morgan & Beatty 2006; see also Chapter 4). Application of supplementary scoring ranked this structure as the number one priority for remediation/removal among all barriers assessed in the pilot study. This indicates that its associated weir pool has only moderate conservation value and complete removal of the barrier warrants consideration. 73

However, this is extremely unlikely given the valuable function (stream flow gauging) provided by the barrier, coupled with the significant investment of capital that went into the construction of the fishway.

The enhancement of native fish migrations through the removal or remediation of instream barriers can come at the expense of facilitating the spread of invasive species

(Starrs et al. 2015). This scenario was encountered in the current study for an instream barrier in the upper Margaret River catchment (barrier #5; Figure 3.7). This structure would have been an ideal candidate for recommendation for remedial works to improve fish migration if not for its role in blocking the further spread of invasive Gambusia holbrooki in this system (Allen et al. 2015). Its prioritisation score was markedly reduced as a result, and removal or remediation of this barrier would therefore be ill advised.

3.6 Conclusion

This study demonstrated that the prioritisation process that was developed can be effectively and efficiently applied across a range of catchments with varying degrees of river regulation and other anthropogenic disturbances. A combination of desktop review of existing barrier information, landholder engagement, and aerial surveys is an extremely effective approach for the identification and initial assessment phase of the barrier prioritisation process. The calculation of prioritisation scores, while using an incomplete dataset, resulted in all of the visually obvious higher priority barriers being ranked as such. Therefore, it proved to be a useful trial of this phase of the process.

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Figure 3.7. Crossing Road ford (barrier #5, score = 58, ranked 12th); an instream

barrier that is blocking the further dispersal of invasive Gambusia

holbrooki in the Margaret River.

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As the human population of south-western Australia increases (Mulholland &

Piscicelli 2012; Infrastructure Australia 2015) and the climate becomes increasingly drier

(Hope et al. 2015), the competition between human demands and ecological requirements for fresh water are expected to intensify, further exacerbating the ecological impacts of instream barriers (see Chapters 1 and 2). However, it is expected that by 2060, many large dams will no longer provide a reliable contribution to the region’s water supply due to declining rainfall and surface flows (Water Corporation 2009; SoE 2011; Smith &

Power 2014). Therefore, the cost-efficiency of maintaining water storage dams in this region is likely to reduce, thus increasing the rate at which they become economically redundant. A key knowledge gap that needs to be addressed is accurate economic modelling of instream barrier decommissioning projects. In the past, removal costs have often been overestimated and the costs of retaining barriers underestimated when deciding whether barriers should be removed or retained. The process developed in this study provides a tool to assist environmental managers in these decisions while delivering benefits to migratory fishes and aquatic ecosystems in drying climatic regions.

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Chapter 4

Use of a vertical-slot fishway by a critically endangered migratory

freshwater fish

4.1 Introduction

River regulation provides numerous socioeconomic benefits, however it also has significant negative impacts on aquatic ecosystems (Ward & Stanford 1979; Poff et al.

1997; Bunn & Arthington 2002; Arthington 2012). One of the most significant impacts is the disruption of habitat connectivity for migratory fishes caused by dams and other instream barriers. In North America, for instance, dams have cut off access to an estimated 40-80% of spawning habitats for anadromous salmonids, contributing to significant declines of a number of species (WWF 2001; Gregory et al. 2002; Sheer &

Steel 2006). The proliferation of instream barriers has also been implicated in the dramatic decline of native fish species in south-eastern Australia including Australian

Bass (Macquaria novemaculeata), Macquarie Perch (Macquaria australasica), Golden

Perch (Macquaria ambigua), and Australian Grayling (Prototroctes maraena) (Harris

1984; Lake & Marchant 1990; Backhouse et al. 2008). Instream barriers are listed as a key threatening process under Australian legislation (Environment Protection and

Biodiversity Conservation Act 1999 (EPBC Act 1999)) and globally recognised as a direct threat to numerous species listed on the International Union for the Conservation of

Nature’s Red List of Threatened Species (IUCN 2015).

One of the most common approaches used to mitigate the impacts of barriers on fish migration is the installation of fishways. Several fishway designs have been 77 developed to accommodate variation in swimming performance of affected fish species and the physical/hydraulic characteristics of different barrier types (reviewed by

Katopodis 1992; Clay 1995; Odeh 1999). The majority of designs (e.g. vertical-slot,

Denil, rock-ramp, bypass) provide a low-gradient channel that abates flow velocity, allowing fish to overcome the vertical drop, or head loss, of an instream barrier. Fishways can be designed to allow multiple species to bypass barriers (e.g. Barrett & Mallen-

Cooper 2006; Calles & Greenberg 2007; Stuart et al. 2008), but often they are designed to provide passage for only a select number of economically important and/or threatened species (e.g. Clay 1995; Naughton et al. 2007).

The Goodga River Fishway in south-western Australia is an example of a structure that was installed to benefit a highly threatened population of Galaxias truttaceus (Morgan & Beatty 2006). This species has a southern Australian distribution

(Allen et al. 2002) with isolated relict populations in the Goodga, Angove and Kent rivers on Western Australia’s south coast (Morgan 2003; Colman 2010; Morgan et al. 2011).

Historically, it was also known from the King and Kalgan rivers, but its absence from those systems in recent surveys suggests that these populations have been extirpated (P.

Close, unpublished data). Presently, the Western Australian population of G. truttaceus is one of only three Australian freshwater fish taxa listed as Critically Endangered under the

EPBC Act 1999.

Most populations of G. truttaceus in eastern Australia are diadromous; however,

Western Australian populations occupy river systems that are essentially land-locked (i.e. they do not reliably discharge to the sea) and Morgan (2003) demonstrated that the

Goodga River population is potamodromous; undertaking an upstream spawning migration in mid- to late autumn. The cue for spawning was hypothesised to be a combination of decreasing water temperature and increasing flow associated with late

78 autumn rainfall (Morgan 2003). The timing of spawning is earlier than most native freshwater fish species in south-western Australia, which typically spawn in winter or spring when flow is near its annual maximum (e.g. Pen & Potter 1991a, b; Pen et al.

1991; Morgan et al. 1995, 2000). After hatching, larvae drift downstream and spend their initial months in the fresh waters of Moates Lake, before migrating upstream into creeks and rivers as juveniles in spring (Morgan 2003; Morgan & Beatty 2006).

Distributional surveys conducted during the 1990s revealed that a low-head concrete weir was restricting the G. truttaceus population in the Goodga River to ~2 km of stream below the weir and blocking access to a further ~2 km of upstream habitat, as well as increasing the risk of avian predation on fish congregating below it (Morgan et al.

1998; Morgan 2003). The weir was, therefore, identified as a significant threat to the population (Morgan 2003), and a vertical-slot fishway was installed in 2003 in order to ameliorate its impacts (Morgan & Beatty 2006). Close et al. (2014) recently demonstrated that juvenile G. truttaceus were able to “climb” the concrete weir structure

(thus bypassing the fishway), but evidently, the numbers of fish ascending in this manner were insufficient to establish a detectable population above the weir prior to the construction of the fishway.

Morgan & Beatty (2006) documented the rapid establishment of G. truttaceus upstream of the weir and also monitored fishway use seasonally for two years after construction of the fishway. Their sampling method involved blocking the exit (i.e. uppermost) slot with a mesh screen and deploying funnel traps in different sections of the fishway to determine passage success. This provided valuable population-level data on seasonal patterns of fishway utilisation, but left many unresolved questions pertaining to utilisation patterns at the individual level and at finer temporal scales in relation to hydrology and other factors. 79

Biotelemetry using passive integrated transponder (PIT) technology has become a common tool in fishway research in recent decades (e.g. Castro-Santos et al. 1996;

Gibbons & Andrews 2004; Stuart et al. 2008; Baumgartner et al. 2010) as it allows movements of PIT-tagged individuals through fishways to be logged remotely, autonomously and continuously. One of the major shortcomings of PIT technology (i.e. short detection ranges) is circumvented by the passage of fish through confined spaces within the fishway, therefore this technology is ideally suited to the study of fishways

(Castro-Santos et al. 1996).

Arguably, the most serious long-term threat to the sustainability of Western

Australian populations of G. truttaceus and other fishes native to the region is the impact of anthropogenic climate change. Temperatures across the region have increased

(Suppiah et al. 2007; IOCI 2008), while rainfall and particularly surface flows have declined sharply since the mid-1970s (Silberstein et al. 2012), and Global Climatic

Models unanimously project a continuation of these trends (CSIRO 2009a, b; Barron et al. 2012; Silberstein et al. 2012; Hope et al. 2015). Stream flow is a cue for the onset of spawning and directly influences the quantity and quality of ephemeral nursery habitats for a number of potamodromous fishes in the region (Beatty et al. 2014a). Therefore, projected flow reductions are likely to have a profound impact on these species.

Similarly, spawning migration in the Goodga River population of G. truttaceus is cued by seasonal shifts in flow and water temperature (sensu Morgan 2003); therefore climate change (particularly declining flow) is likely to have a significant impact on this species and could exacerbate the impacts of instream barriers on migration. However, fine scale

(temporal and spatial) migration data for the species have not yet been gathered and would enable a better understanding of the cues for migration and the potential impact that flow declines will have on its life history.

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The current study aimed to address these questions through the use of biotelemetry. Specifically, this study aimed to gather high resolution temporal and spatial telemetry data to analyse the patterns of utilisation of the Goodga River Fishway by individual G. truttaceus during the annual breeding period. Furthermore, the study aimed to elucidate the relationship between the movement of fish through the fishway and various environmental factors (stream flow, water temperature, diel rhythms). It was hypothesised that movement would be related to increasing stream flow and declining water temperature (sensu Morgan 2003) and that a high proportion of the PIT-tagged adult G. truttaceus would ascend the fishway during the breeding period given previous reports of this species congregating below the weir in autumn (Morgan 2003; Morgan &

Beatty 2006). The information gathered in the current study is expected to provide valuable insight into how projected flow declines due to climate change may alter the migration patterns of this species. These hydro-ecological relationships are extremely valuable in understanding the impacts of instream barriers and, specifically, the functionality of existing fishways, and can be directly incorporated into the barrier prioritisation process outlined in Chapter 3.

A final aim of the study was to estimate the size of the breeding population of G. truttaceus in the Goodga River using recapture data from PIT-tagged individuals. It was hypothesised that the population size would have increased substantially cf. estimates made prior to the construction of the fishway in 2003 (see DoE 2015), given the effective doubling of the available stream habitat (Morgan & Beatty 2006). The results gained from this study will provide valuable data to inform the ongoing management of the

Goodga River Fishway, and contribute to the conservation of this highly imperilled population.

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4.2 Methods

4.2.1 Site description

The Goodga River is located 20 km east of Albany on Western Australia’s southern coast and has a catchment of approximately 60 km2 (Figure 4.1) comprising roughly equal proportions of remnant native vegetation and agricultural land that is predominantly used for cattle grazing and tree plantations. The total stream length of the Goodga River is ca

35 km (including tributaries), approximately half of which flows permanently, due to groundwater input. The river discharges into Moates Lake, a permanent freshwater lake with seasonal connections to nearby Gardner Lake via an ephemeral wetland (Figure 4.1).

The Mediterranean climate of the region features hot, mostly dry summers and cool, wet winters. Mean annual precipitation at the Goodga Weir (1972-2013) is 770.8 mm, with most rainfall occurring between June and August each year (DoW 2015). Mean annual discharge (1964-2012) is 3,865 ML, and follows a similar seasonal pattern with the highest mean monthly discharge occurring in June and the lowest in February (DoW

2015). Water temperature in the Goodga River ranges from 11.7°C in winter to around

23°C in March (Morgan 2003).

Numerous farm dams are found in the lower order tributaries, but no data exist on the proportion of annual discharge that is exploited through water abstraction in the system. The only significant instream barrier impacting the permanent section of the river is the Goodga Weir, a 1.5 m high v-notch structure located ca 2 km upstream of the entrance to Moates Lake (Figure 4.2). The weir has been used for stream flow gauging since 1964.

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Figure 4.1. Map of the study area, showing the location of the Goodga River, Moates

Lake, Goodga Weir and vertical-slot fishway and adjacent hydrographic

features.

83

Figure 4.2. The Goodga Weir and vertical-slot fishway, with Department of Water

(Government of Western Australia) stream gauging telemetry station

pictured at the top right (Photo courtesy of S. Visser).

84

4.2.2 Goodga River Fishway

The Goodga River Fishway is a vertical-slot structure consisting of four sections, each comprising 6-10 cells (0.9 x 0.9 x 0.9 m) separated by aluminium flow baffles with 0.05 m wide slots, and resting pools (2.1 x 1.0 x 0.9 m) at the terminus of each section (Figure

4.2; Marsden 2002). The fishway has a 1:20 gradient with a vertical rise of 1.5 m

(Marsden 2002) and flow velocity in the vertical slots ranges between 0.2 and 1.1 m.sec-1

(Morgan & Beatty 2004). Fishes are able to access the fishway entrance at most tail-water levels except during flood events; however, it is unlikely that fish utilise the fishway during flood pulses as they would be able to swim over or around the weir at such times.

Only five such flood events have occurred since the construction of the fishway in 2003

(DoW 2015).

4.2.3 PIT tags and detection system

Fish were captured from the Goodga River on three separate occasions between May

2013 and March 2015 using either a 5 m long seine net constructed from 1 mm woven mesh and fishing to a depth of 1.8 m, or a fyke net constructed from 2 mm woven mesh and consisting of two 5 m wings, a 1.2 x 0.8 m opening, and a 5 m long pocket with two non-return funnels. Only adult fish ≥80 mm TL (i.e. wet weight ca 2.5 g) were PIT- tagged, following the recommendations of Winter (1996) that transmitter tags should not weigh more than 2% of a fish’s weight out of water. The tagging method used in this study followed that recommended by the tag manufacturer (Biomark 2011). Fish selected for PIT-tagging were kept in an aerated tank containing a solution of AQUI-S® at a concentration of 30 mL.L-1 until they displayed loss of equilibrium. Each individual was then measured for TL to the nearest mm and implanted with a Biomark MiniHPT8™ PIT 85 tag (8.4 x 1.4 mm, 134.2 kHz, 0.033 g) in the posterior section of the peritoneal cavity using a Biomark MK165™ implanter syringe fitted with a 14 gauge needle. The tag insertion point was located on the right hand side of the body, just anterior to the anus

(Figure 4.3). All PIT tags and surgical apparatus were immersed in antiseptic liquid

(Betadine) prior to and after tagging each fish. Tagged fish were immediately placed into an aerated tank and monitored for recovery of equilibrium. Fish that recovered equilibrium were held overnight prior to their release at their original site of capture either upstream or downstream of the fishway. Any fish that did not recover were euthanised in an AQUI-S® solution and preserved in 100% ethanol.

A PIT tag detection system was installed at the Goodga River Fishway on 20th

February 2015. The system comprised two antenna loop assemblies (900 mm x 960 mm) that were custom-built to fit inside the entrance and exit cells of the fishway (Figure 4.4).

The antenna assemblies and cables were fastened to the concrete fishway structure using galvanised saddle clamps and masonry screws. The antenna cables were routed to a secure enclosure on the river bank that housed two Biomark IS1001™ 24 V control nodes, each fitted with a data logger board. Data were logged using 32 GB capacity USB memory sticks and downloaded manually at regular intervals throughout the study. The system was powered by a 0.84 kW solar panel array and 24 V battery bank (Figure 4.4).

Once installed, the PIT tag detection system was range tested using a test tag that was held stationary at increasing distances from all four sides of each antenna loop until it was no longer detected. The average detection range was 33 cm and 30 cm for the fishway entrance and exit antennae, respectively, and the configuration in which both antennae were mounted ensured that tagged fish entering or exiting the fishway would pass within detection range.

86

Figure 4.3. A PIT-tagged Galaxias truttaceus (97 mm TL); point of tag insertion

indicated by yellow arrow.

Figure 4.4. PIT reader antenna installed in the a) entrance and b) exit of the Goodga

River Fishway; c) Biomark IS1001™ 24 V control nodes, data loggers, and 24 V battery bank installed in a secure box at the field site; and, d) 0.84 kW solar panel array used to power the PIT tag detection system. 87

The Data Analysis Toolpak in Microsoft Excel v.14 was used to perform all statistical analyses on data gathered during this study.

4.2.4 Population size estimation

A total of 14 recapture events took place commencing on April 9th 2015 (28 d after the final PIT tagging event) which allowed adequate time for marked individuals to mix with unmarked fish in the wild. On each sampling occasion, fyke nets were set overnight at two fixed sites located 0.7 rkm above and 1.0 rkm below the fishway (Figure 4.1). All captured individuals were counted and scanned using a Biomark 601™ handheld PIT tag reader and the codes of any recaptured individuals recorded. All fish were released at the site of capture. The final recapture event took place on July 13th 2015 (i.e. 95 d after the first recapture event).

Several open population models were explored in the MARK software program; however, the relatively low numbers of recaptures provided inconclusive results. Two factors prompted an exploration of closed population models to estimate the adult population of G. truttaceus in the Goodga River. Firstly, this component of the population is known to occur only in riverine habitats such as those sampled during the recapture events, and secondly, opportunistic sampling ca 2.5 km upstream of the study reach on five separate occasions during the study failed to record any marked individuals, further supporting the assumption of population closure. Nonetheless, as this assumption did not strictly apply in this study, there is a risk that closed population modelling could over-estimate the population due to the possibility of emigration of marked individuals from the survey reach. Despite these limitations, the Lincoln-Petersen Closed population model (using Chapman’s modification) was chosen, given the low recapture rate

88

(Lockwood & Schneider 2000). This model collapsed all recapture periods into a single estimate (a pseudo weighted-average), using the formula:

11 1 1

where N is the estimate of the population abundance, M is the number of fish initially marked, C is the sum of all captured during the recapture events, and R is the sum of the number of fish recaptured during the recapture events. The variance (varN) of the population estimate was then calculated using the formula:

11 var 1 1 R2

Assuming a normal distribution, the confidence intervals were calculated using the formula:

1.96√

4.2.5 Environmental variables

Data on daily stream discharge, stage height, water temperature and rainfall measured at the Goodga Weir gauging station and rainfall gauge during the study period were obtained from the Department of Water, Government of Western Australia. 89

4.3 Results

4.3.1 Detections of PIT-tagged fish

In total, 159 fish (144 G. truttaceus and 15 G. maculatus) were PIT-tagged during the study (Table 4.1), of which ~80% were captured immediately below the Goodga Weir and the remainder were captured 750 m upstream of the weir. A total of 25 individuals

(~16% of tagged fish) were detected by PIT tag readers at least once during the study period (Table 4.2) with just over 50% of these individuals detected on multiple occasions.

Only a single PIT-tagged individual G. maculatus was detected during the study.

4.3.2 Fishway movement patterns

In total, 16 PIT-tagged individuals were recorded successfully passing through the fishway in at least one direction (Table 4.2). Fifteen individuals (all G. truttaceus) successfully ascended the fishway (Table 4.2), two of which ascended on more than one occasion. The maximum number of recorded ascents by an individual fish during the study was three. Of the 15 individuals that successfully ascended, nine were also recorded descending the fishway (Table 4.2), with one fish descending and ascending twice. For individuals that were recorded moving through the fishway in both directions, the lag period between consecutive passage events ranged from 1 to 116 d (median = 7 d). The only individual that descended the fishway without also making a prior or subsequent ascent was the sole representative of G. maculatus that was detected during the study

(Table 4.2). Two of the PIT-tagged G. truttaceus also apparently survived falling over the weir, as they were detected at the fishway entrance subsequent to a logged ascent of the fishway without first having descended the fishway.

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Table 4.1. Summary of PIT tags implanted into fish in the Goodga River, Western

Australia, during the study period.

Tagging date Species n Size range (mm TL)

25/05/2014 Galaxias truttaceus 24 93-149

18/02/2015 Galaxias truttaceus 18 80-114

12/03/2015 Galaxias truttaceus 102 80-167

Galaxias maculatus 15 81-98

Table 4.2. Summary of detections and fishway movement behaviours for PIT-tagged fish in the Goodga River, Western Australia.

Movement

Species Number Number % Ascent Descent Both No tagged (n) detected (d) detected only only directions passage

Galaxias truttaceus 144 24 16.7 6 0 9 9

Galaxias maculatus 15 1 6.7 0 1 0 0

Total 159 25 15.7 6 1 9 9 91

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The mean duration of fishway ascent (57.44 ± 11.50 min) was significantly longer than the mean duration of descent (22.70 ± 3.04 min) t(20) = 2.8480, p < 0.01. There was no significant correlation between fish length (TL) and duration of fishway ascent (r(16)

= 0.1799, p = 0.4751) and the sample size was insufficient to reliably test for a correlation with duration of fishway descent. A clear trend was evident; however, in the diel timing of fishway movements, with 85.2% of passage events taking place during nocturnal hours. Only four passage events took place at other times of the day; three (two ascents and one descent) during crepuscular periods and one during daylight hours. The frequency of PIT detections in the fishway was much higher during nocturnal periods

(74.6%) compared to diurnal periods (25.4%) (Figure 4.5).

Nine individuals were detected at one of the gates (all but one at the fishway entrance) without subsequently passing through the fishway (Table 4.2). For fish detected at the fishway entrance (excluding those that had just completed a descent of the fishway), 71.9% of the detection events (n=64) did not result in an ascent of the fishway.

Four fish were detected at the entrance of the fishway on multiple days, with one of these fish detected on 25 different dates over a 59 d period between March and May including a run of detections over 11 consecutive days (March 13-23).

4.3.3 Influence of environmental variables on fishway movements

Successful passage through the fishway mostly coincided with flow pulses (Figure 4.6).

Early in the study period, daily discharge was low and steady at <3 ML.d-1, punctuated by occasional minor flow pulses between 3 and 4 ML.d-1 and successful ascents took place on two of these flow pulses (Figure 4.6). A flood pulse (approximately an order of magnitude larger than the second highest pulse recorded during the study period) 93

12

10

8

6

Frequency (%) Frequency 4

2

0 01234567891011121314151617181920212223 Hour

Figure 4.5. Frequency of PIT detections of tagged G. truttaceus (white bars) and G.

maculatus (red bar) in the Goodga River Fishway during different hours of

the day. Grey areas indicate night periods based on the local sunrise and

sunset times at the mid-point of the study.

94

5 120 Ascents Descents 80 4 40

20 3

15

2

10 Daily Discharge (ML)

1 * Number of Successful Passages 5

0 0

21/2 7/3/15 21/3/1 4 1 2 16 30 13 /4/15 8/4/1 / 5 / / / /1 5 5 6 /15 5 /1 /15 /15 5 5 5

Figure 4.6. Successful fishway passage events (bars) recorded for PIT-tagged fish in

the Goodga River, Western Australia, during autumn 2015 (* solitary G.

maculatus passage). Total daily discharge (ML) is shown as a line graph.

95 occurred on April 10th but notably, no fish passages coincided with this event (Figure

4.6). The second half of the study period was characterised by more frequent flow pulses

(discharge up to ~11 ML.d-1), and the majority of successful fish passages took place during this part of the study, 68.4% of which coincided with a flow pulse event (Figure

4.6). The highest number of successful fish passages was recorded in conjunction with the first flow pulse (i.e. May 4th) that came after the April 10th flood pulse; six ascents and two descents were recorded at this time (Figure 4.6). During the autumn breeding period on days when successful passage events were recorded, the number of passage events per 24 h period was positively correlated with the change (flux) in daily discharge in the antecedent 24 h period (r(12) = 0.8170, p < 0.001) (Figure 4.7).

4.3.4 Population size estimate

In order to minimise time in captivity and handling stress, fish sampled during the recapture phase of the study were counted but not individually measured. However, this led to some imprecision in the population estimate as the exact number of individuals >80 mm TL (i.e. the size class that was PIT-tagged) was not recorded. The mean estimate of three observers of the proportion of individuals captured that were > 80 mm TL (i.e.

~70%) was therefore used to generate the total count of recaptured individuals (C). The population estimate (N) of individual G. truttaceus > 80 mm TL in the study reach generated by the model was 13,910 (lower/upper 95% CI = 8,948-18,873). A simple extrapolation of this result yielded an estimate of ~ 20,000 individuals of all size classes

(excluding larvae) in this section of the river; however, this may have been an overestimate (see section 4.4.4 for further discussion).

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Figure 4.7. Scatterplot of the number of successful fishway passage events occurring

in a 24 h period versus the change in daily discharge in the antecedent 24 h

period. Days when no successful passage events were recorded are not

shown. Red circle indicates the solitary passage event recorded by a G.

maculatus.

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4.4 Discussion

4.4.1 Fishway passage and flow

The current study confirmed the hypothesis that movement through the fishway by G. truttaceus during the annual spawning period is related to streamflow; as revealed by a significant positive correlation between river discharge and successful fishway passage.

This supports the findings of Morgan (2003), who reported that increasing water level

(i.e. flow) and concomitant decreasing water temperature were cues for the commencement of the annual spawning migration for the species in this system.

Similarly, Tasmanian populations of G. truttaceus undertake spawning migrations during periods of stream flow (Humphries 1989) as do numerous other species of Australian galaxiids (e.g. Pollard 1971; Hopkins 1979; Mitchell & Penlington 1982; Pen & Potter

1991b; Beatty et al. 2014a).

A strong correlation was observed between the flux in daily river discharge and successful fish passage for the majority of flow levels recorded during the study, but notably, no passage events were recorded during the highest magnitude flow of the study

(110.731 ML in 24 hr), a flood pulse in the 99th percentile for daily discharge in the system. It is unlikely that this observation was attributable to hydraulic conditions within the fishway itself, as vertical-slot fishways are designed to provide stable flow conditions irrespective of prevailing river levels (Larinier 2002). This observation may simply reflect a behavioural preference of this species to defer fishway passage attempts during flood events. Alternatively, the high tailwater levels may have permitted fish migrating upstream to bypass the fishway completely, thus evading detection. Yet another possible explanation is that excessive turbulence in the tailwater during this flood pulse may have distracted or disorientated migrating fish and prevented them from locating or accessing

98 the narrow (i.e. 0.05 m wide) fishway entrance, as has been hypothesised in fishway passage studies elsewhere in high flow conditions (e.g. Barry & Kynard 1986; Bunt et al.

1999). Mallen-Cooper (1996) reported a correlation between fishway movements and minor flow events for two percichthyids in the Murray River, whilst other studies have demonstrated that periods of low flow are usually more important for fish migration than high flow periods, with a higher abundance and diversity of fishes using fishways during low flows (Stuart & Berghuis 2002; Stuart & Mallen-Cooper 1999).

4.4.2 Diel trends in fishway utilisation

Fishes commonly display diel patterns of fishway utilisation that mirror patterns of regular activity, i.e. nocturnally active species generally utilise fishways at night, and vice versa (sensu Thiem et al. 2012). The clear preference for nocturnal fishway movement detected during the current study therefore suggests that G. truttaceus is probably a nocturnally active species. Other galaxiid species are known to be nocturnally active

(Glova & Sagar 1989; Rowe & Chisnall 1996; David & Closs 2003; Hardie et al. 2006); however, Morgan and Beatty (2006) detected no significant difference between the total numbers of G. truttaceus using the Goodga River Fishway at night versus during the day over two consecutive autumn seasons. They did, however, find that fish using the fishway at night were significantly larger than those using it during the day (Morgan & Beatty

2006). Unfortunately, the current study was unable to confirm this ontogenetic shift in the diel pattern of fishway utilisation due to the size limitation of fish that could be implanted with PIT tags. However, the results of the current study clearly supported the findings of

Morgan and Beatty (2006) that larger fish display a marked preference for nocturnal fishway movement, at least during the autumn breeding season. 99

The explanations for this strong nocturnal bias among the adult population are unclear, although it may reflect a strategy to minimise the risk of predation. Keefer et al.

(2013) hypothesised that lampreys in a North American river used fishways at night to reduce potential exposure to diurnally active salmonid predators. The Goodga River lacks piscean predators; however, avian predators such as the White-faced Heron (Egretta novaehollandiae) have previously been identified as a threat to the population, particularly during the annual breeding migration when fish congregate below the

Goodga Weir in high abundance (Morgan 2003). Piscivorous waterbirds are typically diurnal feeders that rely on sight to stalk or ambush prey (Marchant & Higgins 1990) and the avoidance of such predators may account for the bias towards nocturnal utilisation of the fishway by adult G. truttaceus. The prevalence of smaller individuals using the fishway during daylight hours (Morgan & Beatty 2006) remains unexplained.

4.4.3 Site fidelity and fishway residency

The results of the current study suggest that there is a degree of site fidelity amongst mature age classes during the breeding season with several individuals recaptured at the same site on multiple occasions throughout the study. Furthermore, the hypothesis that a substantial proportion of the PIT-tagged adult population would ascend the fishway during the breeding period was rejected, as only ~8% of tagged individuals originally captured below the weir were detected ascending the fishway. These results show that upstream migration for breeding is not a behaviour that is universal across the adult population. Whilst suitable spawning and nursery habitat does exist above the weir, as evidenced by the capture of abundant larvae drifting downstream at the Goodga Weir in winter (P. Close, unpublished data), the fact that the majority of PIT-tagged individuals

100 did not ascend the fishway during the annual breeding period, and that ~23% of those tagged and released above the fishway actually descended the fishway at this time, suggests that these critical habitat types may be more prevalent downstream of the weir.

That the population remained extant following construction of the weir in the 1960s, is clear evidence that the lower reaches of the river contain the requisite habitat types to sustain this population, albeit in a spatially confined range. In an anthropogenically impacted system such as the Goodga River, the presence of suitable spawning habitat both upstream and downstream of the weir and the maintenance of fishway functionality are factors that favour the persistence of the population.

Four individuals were recorded visiting the fishway entrance on multiple occasions, providing further evidence of site fidelity amongst the adult population of G. truttaceus in this system. One of these fish was detected at the fishway entrance on 25 separate dates over a 59 d period but did not pass through the fishway once during this time. The time spent within detection range of the entrance gate varied from 52 s to over

24 h and there was a degree of regularity about the time of initial detection. On just over a third of these dates, initial detection occurred during the pre-dawn period (4:00-6:30

AM). This pattern of visitation suggests that the fishway may have been used as a regular habitat by this individual. A second individual also showed a similar pattern of fishway visitation over an 18 d period between June and July. This fish was detected at the fishway entrance on seven separate occasions and on over two thirds of these occasions the time of initial detection also occurred during the pre-dawn period (6:30-7:30 AM).

These data could be interpreted as evidence of home ranging behaviour by these individuals. The explanation for this pattern of visitation to the fishway by these individuals is unclear but may have been for the purpose of feeding, accessing shelter, social interaction, or perhaps even spawning. In previous fish monitoring studies of the 101

Goodga River Fishway, the capture of so called “resident” fish that seemingly used the fishway as regular habitat was common (D. Morgan, pers. comm.). Although infrequently reported in the scientific literature, at least one other study has documented fishway residency, for White Sturgeon (Acipenser transmontanus) in a fishway on the Columbia

River (USA) for a period of 6 months (Parsley et al. 2007).

4.4.4 Limitations of the study

One of the main shortcomings of the PIT tag reader array used in this study was the use of only two antennae (due to budgetary restrictions), as this precluded the determination of passage efficiency (i.e. the percentage of successful attempts at negotiating passage through the entire length of the fishway). Other fishway studies have employed arrays consisting of four or more antennae deployed at regular intervals along the fishway to achieve this objective (e.g. Baumgartner et al. 2010; Stuart et al. 2008). In this study, there were numerous instances of fish being detected at either the entrance or the exit gate but not passing through the fishway. The detection logs revealed varying periods of time spent at the gate with a single or multiple gaps in detections that ranged from seconds to tens of minutes. Unfortunately, owing to the configuration of the two antennae array, it was not possible to ascertain the direction of movement (either into the fishway or out into the river) of fish during these detection gaps. However, even if it was known that the fish had made an attempt to move through the fishway only to return to its starting point, it is not possible to know for certain if this assumed “passage failure” was the result of the individual’s incapacity to successfully negotiate the fishway or simply a behavioural preference to temporarily utilise the fishway structure as habitat.

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The assumption that detected individuals not passing completely through a fishway have “failed” to pass potentially obfuscates studies on fishway passage efficiency. To highlight this point, early in the current study, two individual fish were detected at the fishway entrance and did not pass through the fishway. However, later in the study period, these same individuals successfully ascended the fishway. It is unlikely that flow conditions accounted for the earlier “failed passage” as flow velocity remained relatively stable within the fishway throughout the entire study period. Rather, this likely represented a behavioural preference on the part of these fish to not pass through the fishway at the earlier time.

Whilst the majority of detected individuals in the current study were captured and tagged in February/March 2015, two fish were detected that had been tagged in 2014, almost a full year prior to the installation of the PIT tag readers at the fishway. This indicates that G. truttaceus is capable of longer-term survival post-tagging. One of these fish was 149 mm TL and therefore likely to be in its 7th or possibly 8th year of life at the time of initial capture, based on the von Bertalanffy growth curve data for the Goodga

River population calculated by Morgan (2003). Laboratory trials were not conducted to determine the likelihood of tag retention or the rate of post-tagging mortality in G. truttaceus. Other studies of juvenile salmonids of a similar size range to the fish tagged in this study, have recorded high rates (typically >99.99%) of tag retention and low incidences (generally <1%) of post-tagging mortality (e.g. Dare 2003; Larsen et al. 2013;

Huusko et al. 2016). In the current study, mortality was recorded in 3.1% of fish within the initial 12 h holding period post-tagging; however, tagged fish that survived the holding period showed no apparent adverse effects of the procedure and the survival rate, based on the results of other studies, was expected to be high. Nevertheless, some additional fish may have suffered delayed mortality during the study period due to 103 tagging or for other reasons (e.g. predation), or may have shed their tags, thus potentially resulting in an overestimation of population size and underestimation of the incidence of fishway utilisation for G. truttaceus in the Goodga River.

4.4.5 Management implications

This study is one of only a few to estimate population numbers of a native Australian freshwater fish in the wild using mark-recapture methods and is the first to use PIT tags for this purpose. The population size estimate for G. truttaceus in the Goodga River in this study was an order of magnitude greater than that estimated prior to the construction of the fishway in 2003 (see DoE 2015), which supports the hypothesis that the fishway has led to an increase in numbers. However, the method used to estimate the population in the earlier study is unclear, with the text stating that it was “based on sampling surveys” (DoE 2015), whilst that of the current study was based on mark-recapture data albeit with some shortcomings (see above section 4.4.4), therefore a degree of caution is needed when comparing these results. Nevertheless, it is not unreasonable to surmise that the population of G. truttaceus in the Goodga River has increased substantially as a result of the effective doubling of the available habitat following construction of the fishway.

As a large proportion of the Goodga system upstream of the fishway is impacted by cattle grazing, the potential exists for additional population growth should stream restoration initiatives such as fencing and revegetation of degraded riparian zones be undertaken in this catchment in the future.

This study showed that successful passage through the fishway was strongly linked to flow which presents challenges for maintaining functionality of the fishway in a changing climate. However, projecting the impacts of climate change on breeding and

104 recruitment of this species is difficult. Global Climate Models (GCMs) unanimously project a continuation of the drying climatic trend that has prevailed in south-western

Australia since the mid-1970s (Silberstein et al. 2012; Smith & Power 2014). The scale of climatic drying; however, is projected to vary across the region (DoW 2010;

Silberstein et al. 2012). The south coast, which is home to all Western Australian populations of G. truttaceus, is projected to experience only minor declines in rainfall and stream flow; however, significant seasonal declines have been forecast to occur in summer and autumn (DoW 2010). This study showed that fishway movement in autumn may only be of minor importance to this population; therefore projected flow declines may have a negligible impact on spawning migrations. However, reduced rainfall and flows during the autumn breeding season may have broader impacts in catchments housing this species by reducing both the quantity and quality of available spawning habitat, which in turn may lead to a decline in recruitment levels.

Data from the G. truttaceus population in the , another system that is impacted by instream barriers and water abstraction, also show a delay of approximately one month in the timing of spawning and recruitment cf. the Goodga population (P. Close, unpublished data). Moreover, during the study period, a substantial proportion of adult G. truttaceus in the Goodga River population were observed to maintain a “running ripe” reproductive condition for an extended period of time

(approximately one month), presumably in response to a delay in the onset of optimal spawning conditions (P. Close, unpublished data). Additionally, allopatric populations of

G. truttaceus in Tasmania have been shown to spawn at different times of year

(Humphries 1989), as have discrete populations of G. maculatus on Western Australia’s southern coast (Chapman et al. 2006). There certainly appears to be a degree of biological plasticity in terms of the timing of reproduction for G. truttaceus and other 105 galaxiids, which may offer a selective advantage in coping with changing climatic conditions; nonetheless, reductions in amount of aquatic habitat or flow periods may increase inter- and intra-specific competition for resources, and in turn, may cause a decline in the carrying capacity of these systems.

Perhaps the most significant finding of this study was the low percentage of tagged adult fish that utilised the fishway during the annual breeding period. It is reasonable to infer that the majority of fish in the study reach spawned at sites located downstream of the Goodga Weir, but for a minority of the breeding population at least, the fishway presumably serves a useful function in facilitating access to habitats upstream of the Goodga Weir. Morgan and Beatty (2006) documented large numbers of young-of- year G. truttaceus migrating upstream through the fishway in late spring (mean >1100 individuals/24 hr period), therefore the fishway probably provides a more important ecological role in facilitating juvenile recruitment and dispersal each spring, rather than as a migratory pathway for spawning adults in autumn.

The data gathered during this study are useful in the application of the barrier prioritisation process (see Chapter 3), in that they highlight the complexity of the impacts that instream barriers can have on migratory fish species. In this instance, the Goodga

Weir is probably not a significant barrier to adult breeding migration given the low incidence of upstream migration of PIT-tagged adults through the fishway; however, its overall impact on limiting the range (and size) of the population prior to its remediation with a vertical-slot fishway showed the significant impact that it had on restricting migration of juveniles of this species. The clear benefit delivered to the Goodga River population of G. truttaceus through the remediation of the Goodga Weir should serve as a template of similar conservation outcomes that could be achieved in the nearby Angove

River which also houses a population of G. truttaceus that is impacted by two significant

106 instream barriers (i.e. a v-notch gauging weir and a low-head dam used to supply potable water to the nearby town of Albany).

107

Chapter 5

The biology and migration patterns of the Little Pygmy Perch,

Nannoperca pygmaea (Teleostei: Percichthyidae), including comparisons

with two sympatric percichthyids

5.1 Introduction

Potamodromous fishes migrate exclusively within fresh waters, usually for the purpose of breeding, and complete their entire life-cycle within fresh waters (Myers 1949). The evolution of this mode of behaviour across a broad diversity of fishes worldwide is evidence that the selective advantages outweigh the energetic costs and risks inherent to migration (Alerstam et al. 2003; Koehn & Crook 2013). Reproduction is a common driver of potamodromous migrations with propagules or young being released into nursery habitats with characteristics more conducive to larval growth and survival (e.g. fewer predators and/or ecological competitors, higher habitat complexity and food availability) cf. normal adult home ranges, thereby maximising recruitment (Lindsey et al. 1959; Northcote 1984, 1997). Recruitment success may also be enhanced via upstream migration in lotic systems to offset denatant drift of propagules and larvae into unsuitable brackish or saline waters (e.g. Reynolds 1983). Potamodromy can also serve non- breeding purposes, such as in the exploitation of seasonal or stochastic but spatially discrete food resources (i.e. trophic migrations (e.g. Northcote 1997; Douglas et al. 2005;

Sternberg et al. 2008)), the colonisation of ephemeral habitats (e.g. Perry & Bond 2009), or for survival through migration into habitats that offer refuge from adverse environmental conditions such as desiccation or extreme temperature (Northcote 1997;

108

Crook et al. 2010; Koehn & Crook 2013). However, the exact reasons underpinning migration are major knowledge gaps for many potamodromous species (Koehn & Crook

2013).

The discovery of Nannoperca pygmaea in the Hay River system in 2007 resulted in the first description of a new freshwater fish species in south-western Australia for almost 30 years (Morgan et al. 2013). The life-cycle of this rare percichthyid is yet to be determined, but its occurrence in the Mitchell River (Morgan et al. 2013), an ephemeral tributary of the Hay River, suggests that it may be potamodromous. Three other percichthyids occur in freshwaters of south-western Australia, i.e. Bostockia porosa,

Nannatherina balstoni, and the congeneric N. vittata; and all three species have been demonstrated to undertake potamodromous migrations between permanent mainstem refuges and lower order ephemeral tributaries (Pen & Potter 1990; Pen & Potter 1991a;

Beatty et al. 2014a). A reliance on potamodromous migration into ephemeral habitats renders fish taxa highly susceptible to the impacts of altered hydrology caused by anthropogenic stressors such as instream barriers (Lucas & Baras 2001; Gregory et al.

2002) and climate change (Morrongiello et al. 2011; see also Chapter 2). Beatty et al.

(2014a) determined that the number of fish migrating to spawn in ephemeral tributaries of the system (including N. balstoni and N. vittata) was correlated with discharge. Probable reductions in reproduction, recruitment, and populations of potamodromous fishes (Beatty et al. 2014a) wrought by drying climatic conditions in the region (Suppiah et al. 2007; CSIRO 2009; Hope et al. 2015) will present significant conservation challenges in the future, potentially exposing these species to increased risk of range contraction, localised extirpations, or even extinctions.

Prior to this study, the known distribution of N. pygmaea consisted of only three sites within a section of the Hay River system covering 0.06 km2 (Morgan et al. 2013). 109

Extinction risk is particularly high for spatially restricted taxa with small population sizes that are exposed to anthropogenic stressors such as habitat loss, introduced species and climate change (Frankham et al. 2002; O’Grady et al. 2004; Tedesco et al. 2013).

Therefore, this study aimed to determine the key threats to N. pygmaea and elucidate if any significant barriers (either physical or hydrological) existed that could account for its highly restricted known range. The study also aimed to elucidate key aspects of the biology and life history of N. pygmaea in the Hay River, including population size in baseflow refuge habitats and the timing and ecological reasons underpinning migration patterns. Such information can aid the process of assessing the status of threatened species (Sjögren-Gulve & Ebenhard 2000), as well as formulating and prioritising conservation actions (IUCN 2014). It was hypothesised that N. pygmaea would possess similar life-history traits (e.g. reproductive biology, diet, age and growth) to N. vittata given the close phylogenetic relationship between these two sympatric species (Unmack et al. 2011; Morgan et al. 2013).

Given the close physical resemblance of these two species and the superficial resemblance to N. balstoni (Figure 5.1; refer to Morgan et al. 2011 for distinguishing characteristics), it was also hypothesised that N. pygmaea may have potentially been overlooked in previous surveys of other systems. Therefore, another major aim was to determine the extent of its distribution, particularly during the critical summer/autumn baseflow period, by undertaking a comprehensive sampling regime in the Hay River and other nearby systems to search for previously undetected occurrences of this .

Using these data, the study finally aimed to formally assess the conservation status of N. pygmaea, enabling the development of management actions to help conserve the species.

110

Nv

Nv

Nv Np

Nv Nb

Figure 5.1. Three sympatric pygmy perch species (Nv – N. vittata, Np – N. pygmaea,

and Nb – N. balstoni) captured from the lower Mitchell River, south-

western Australia (Photo courtesy of S. Beatty).

111

5.2 Methods

5.2.1 Study area

This study focussed on the Hay River and one of its major tributaries, the Mitchell River

(Figure 5.2), in order to determine key aspects of the biology and life history of N. pygmaea. To determine its distribution, other nearby catchments including the Denmark,

Kent and Bow were also sampled (Figure 5.2). These rivers are located in the southernmost part of Western Australia which experiences a highly seasonal

Mediterranean climate with most rainfall and stream flow occurring during winter and spring (DoW 2015; BoM 2015). The average annual rainfall exceeds 1,000 mm near the coast and diminishes to ~600-700 mm at the headwaters of these catchments (BoM

2015).

The Hay River arises ~60 km inland of the town of Denmark and discharges into the

Wilson Inlet, a major regional estuary (Figure 5.2). The catchment covers an area of

~1,300 km2, 70% of which is cleared for agriculture (sheep, beef, dairy, and vineyards), with the remainder comprising native, dry sclerophyll forest dominated by jarrah

(Eucalyptus marginata) and marri (Corymbia calophylla) (Pen & Apace 1995). The mean annual discharge at the Sunny Glen gauging station on the lower reaches of the Hay River is 50.6 GL (S.E. ± 6.8) (DoW 2015). The Mitchell River is the largest tributary of the

Hay River with a catchment area of ~150 km2 that consists almost entirely of remnant native forest. Its confluence with the Hay River lies some 13 river kilometres (rkm) upstream of the interface of the Hay River and the Wilson Inlet. The mean annual discharge at the Beigpiegup gauging station in the upper reaches of the Mitchell River is

3.1 GL (S.E. ± 0.4) (DoW 2015).

112

Figure 5.2. Overview map of the study area showing the locations of all sites sampled

for Nannoperca pygmaea between 2009 and 2014.

113

The also discharges into the Wilson Inlet (Figure 5.2) and has a catchment area of 671 km2, comprising mostly remnant native vegetation (Ward et al.

2011). Wet sclerophyll forest occurs in the higher rainfall zone near the coast and dry sclerophyll forest further inland. Since the late-1970s, ~18% of the catchment (mostly in the upper reaches) has been converted from cleared farmland to timber plantations as part of a catchment-scale program aimed at reversing the effects of secondary salinisation and recovering the Denmark River as a potable water resource (Schofield et al. 1989; Bartle

1991; Ferdowsian & Greenham 1992; Ward et al. 2011). In 2009, only 7% of the catchment area consisted of cleared farmland (Ward et al. 2011). The mean annual discharge of the Denmark River at the Mt Lindesay gauging station located in the middle reaches of the catchment is 25.5 GL (S.E. ± 2.3) (DoW 2015).

The has a total catchment area of ~2,400 km2 and arises ~80 km inland of the coast north of the town of Denmark and discharges into the Irwin Inlet

(Figure 5.2). The catchment predominantly comprises remnant native vegetation (both wet and dry sclerophyll forests) and timber plantations (combined ~68%) with the remainder of the catchment cleared for agriculture (De Silva et al. 2006). The majority of clearing has taken place in the upper reaches, which has led to secondary salinisation (De

Silva et al. 2006). Between 1990 and 2002, mean annual salinity in the Kent River was

3,180 mg.L-1 in the upper catchment and 1,480 mg.L-1 in the lower catchment (De Silva et al. 2006). Plans are underway to alter existing land and water uses in the Kent catchment in order to recover its surface water resources to potable levels (i.e. ≤ 500 mg.L-1) by 2030 (De Silva et al. 2006). Mean annual discharge at the Styx Junction gauging station in the lower catchment is 76.6 GL (S.E. ± 6.4) (DoW 2015).

114

5.2.2 Distributional sampling and fish monitoring

Prior to this study, N. pygmaea was known from only two sites in the Hay River in the vicinity of the Mitchell River confluence, and one other site in the lowermost reaches of the Mitchell; a combined area of 0.06 km2 (Morgan et al. 2013). To gain a better understanding of the extent of the distribution of this species in the Hay River system, 43 sites throughout the catchment were sampled between 2009 and 2014, with most effort focussing on non-salinised tributaries in the general vicinity of the previously known distribution (Figure 5.2). Fish sampling was initially undertaken at sites in the Mitchell

River and adjacent parts of the Hay River mainstem during winter and spring of 2009, with further exploratory sampling in the Mitchell River during 2010.

As the Hay River only flows intermittently in winter/spring each year and contracts to isolated permanent pools during summer/autumn, one of the major aims of the study was to ascertain the distribution of N. pygmaea in this system during the critical baseflow period. To achieve this, an aerial survey was conducted in March 2013 (see section 3.3.2 for a description of the methods used) to map all permanent aquatic refuge habitat in the segment of the Hay catchment that included all previously known sites of

N. pygmaea. The aerial survey covered the Hay River mainstem between Sunny Glen gauging station (34.91° S, 117.48° E) and farmland located at the boundary of Mt

Lindesay National Park (34.88° S, 117.52° E), as well as the lower reaches of the

Mitchell River (upstream to 34.85° S, 117.44° E). A number of key refuge pools that had been identified during the aerial survey were sampled in April 2013. Sampling effort was widened in winter-spring of 2013 to include several seasonally flowing tributaries of the

Hay River (i.e. Mitchell River, Sheepwash Creek, Sleeman Creek), and in 2014 was further expanded into neighbouring catchments (i.e. Denmark, Kent, Bow) (Figure 5.2). 115

All sites where N. pygmaea occurred were mapped in ArcGIS™ Desktop 10.2 and minimum convex polygons (α-hulls) were drawn to calculate the extent of occurrence

(EOO) (i.e. the area contained within the shortest continuous boundary which can be drawn to encompass all the known, inferred or projected sites of present occurrence of a taxon, excluding cases of vagrancy) following IUCN guidelines (IUCN 2014). Area of occupancy (i.e. the area within the EOO known to be occupied by a taxon, excluding cases of vagrancy) was estimated by overlaying the distribution map with a 1 km2 grid and counting the number of grid squares that contained N. pygmaea sites (IUCN 2014).

Fish were sampled using either seine nets (5 m or 10 m in length) constructed from 1 mm woven mesh and fishing to a depth of 1.8 m, or fyke nets constructed from 2 mm woven mesh and consisting of two 5 m wings, a 1.2 × 0.8 m opening, and a 5 m long pocket with two non-return funnels. Seine net sampling involved three replicate net drags at each site with the two-dimensional area of coverage of each net drag estimated in order to calculate the mean density of fishes (individuals.m-2). Fyke nets were deployed in pairs, with one net facing upstream and the other facing downstream, in order to detect the directionality of fish movements. Nets were set in the afternoon and checked the following morning. The total duration of each fyke net deployment was used to calculate catch per unit effort (CPUE) expressed as individuals.net-1.h-1. Note that at sites where the stretched wings of the fyke net did not block the entire width of the stream channel, numbers of individuals were standardised by multiplying the inverse of the proportion of stream blocked by the net by the number of individuals captured in the net as per Beatty et al. (2014a).

All fish sampled during the study (irrespective of capture method) were identified to species level, counted, measured for total length (TL) to the nearest 1 mm, and examined externally to determine if they were in spawning condition (running ripe). This

116 was ascertained by applying gentle pressure to the abdominal region of each fish and checking for exuded sperm or ova (Morgan & Beatty 2006). All fish were released alive at the site of capture, with the exception of 12 female N. pygmaea in breeding condition that were euthanised in an ice slurry and stored in 100% ethanol for later dissection in order to elucidate life history characteristics (see section 5.2.3 below). The number of N. pygmaea specimens that were retained for dissection in the current study was intentionally low, so as to minimise destructive sampling of this rare species.

5.2.3 Biological characteristics

The gonads of the aforementioned female specimens were fixed in Bouin’s solution for

24 h and transferred into 100% ethanol for long term storage. Two gonads were prepared for histological assessment, firstly by dehydration in a series of alcohols before they were embedded in paraffin wax, sectioned at 6 µm and stained with Mallory’s trichrome. The sections were viewed under a compound microscope and used to describe the composition of the ovaries of reproductively active females.

Five of the twelve remaining gonads were used to estimate fecundity by applying a modification of the methodology of Llewellyn (1979). These gonads came from females that ranged in length between 41 and 51 mm TL. A dissecting microscope fitted with an eyepiece graticule was used to count oocytes that were assigned to arbitrary size categories. The number of large (i.e. >700 µm) oocytes were counted in full for each gonad, whereas numbers of smaller sized oocytes were estimated by excising a portion of the gonad, weighing it to the nearest 1 mg, counting oocytes within the sub-sample and applying the following formula to determine fecundity: Fecundity = N × (W1 / W2); 117

where, N = oocyte count; W1 = the wet weight of the whole gonad, and; W2 = the wet weight of the gonad sub-sample.

Sagittal otoliths were also removed from these 12 specimens, placed in a black dish containing glycerol, and examined under reflected light using a dissecting microscope in order to count the number of hyaline (i.e. translucent) zones, which is often a reliable indicator of the age of a fish (see Bagenal 1973). The limited number of fish otoliths used precluded validation of annual formation of a single hyaline zone; however, it was assumed that the number of hyaline zones would be a reliable indicator of age for this species given that this aging technique has been successfully validated for all other native percichthyids in south-western Australia (e.g. Pen & Potter 1990, 1991a; Morgan et al. 1995). The age estimates obtained from otoliths were used in conjunction with an analysis of cohort progression in the monthly length-frequency histograms for the Hay-

Mitchell sub-population in order to estimate growth rates and longevity.

Following dissection, the intestinal tract was also removed from each of the specimens used in the age and growth study. These were examined under a dissecting microscope in order to provide a qualitative evaluation of the dietary composition of this limited sample of 12 adult female N. pygmaea specimens captured in the Mitchell River

(Hay River system) during August/September 2013.

5.2.4 Environmental variables

Monthly precipitation data for Denmark and Denbarker (located centrally in the study area) were obtained from the Bureau of Meteorology website (BoM 2015).

Physicochemical variables were measured using a YSI™ Professional Plus multimeter

118

(YSI Inc., Yellow Springs, OH, USA). Mean (±1 S.E.) estimates of three replicate measurements of water temperature (°C), pH, dissolved oxygen (%), salinity (‰ NaCl), total dissolved solids (‰), and electrical conductivity (µS.cm-1) were calculated at each sampling site on each sampling occasion.

5.2.5 Visible Implant Elastomer (VIE) tagging and population size estimation

During the baseflow period in 2014, the population sizes of N. pygmaea, N. vittata and N. balstoni were estimated in two isolated refuge pools on the Hay River mainstem (Figure

5.3) using a mark-recapture technique. These pools were chosen as they were the only baseflow refuges for N. pygmaea known to exist prior to this study. Water depth was measured at 15 randomly chosen points in each pool, the co-ordinates of which were logged using a handheld GPS (Garmin eTrex 30) and the bathymetry of the pool was modelled using the IDRISI GIS Analysis software package (Clark Labs, Worcester, MA,

USA) (Figure 5.3). Bathymetry modelling gave an estimate of the surface area and volume of the pools of approximately 491 m2 and 134 m3 for Refuge Pool 1, and 615 m2 and 733 m3 for Refuge Pool 2. The maximum recorded depths were 1.2 m and 3.1 m in

Refuge Pool 1 and 2, respectively.

A total of five fish sampling events took place at approximately one week intervals between March 6th and April 9th 2014. On each sampling occasion, eight double- winged fyke nets (construction of nets as described above) per pool were set in the afternoon at fixed locations and retrieved the following morning. All N. pygmaea and N. balstoni individuals that were captured were marked for this component of the study; however, as the number of N. vittata captured was high, only a subsample of individuals were marked on each occasion. 119

Refuge 1 Refuge 2

Figure 5.3. Photographs and bathymetry models of two baseflow refuge pools in the

Hay River near the confluence with the Mitchell River, Western Australia,

where a mark-recapture program was undertaken to estimate population

sizes of pygmy perch species.

120

Pygmy Perch selected for tagging were placed in a 100 L aerated holding tank and transferred in small batches (5-10 individuals at a time) into an aerated 20 L bucket containing a solution of AQUI-S® at a concentration of 30 mL.L-1 until they displayed loss of equilibrium. Anaesthetised fish were measured for TL to the nearest 1 mm and injected with two strips (~3 mm in length) of visible implant elastomer (VIE) in muscular tissue on the dorsal surface of the body using a manual elastomer injector loaded with a

0.3 cc hypodermic syringe (NMT 2008). VIE was chosen as it has previously been demonstrated to have the desirable combination of a high tag retention rate and high rate of survival in pygmy perches (Price 2009). Each fish was given a dual-tag combination that was preselected from six potential tag locations, three on either side of the body

(Figure 5.4), and five colours (red, green, blue, yellow, orange). This enabled up to 375 individuals of each species to be uniquely identifiable for the duration of the mark- recapture study. Following VIE tagging, fish were placed in an aerated 100 L tank to recover. Upon regaining equilibrium, tagged fish were released at the point of initial capture.

On each sampling occasion after the initial event, all captured pygmy perch were thoroughly inspected for VIE tags, with recaptured individuals identified, measured for

TL to the nearest 1 mm and returned alive to the water. All non-tagged pygmy perch were anaesthetised, tagged and released following the methods just described, except for N. vittata, of which only a subsample was tagged on each sampling occasion, owing to the large numbers of this species that were captured.

Closed population modelling was conducted on the individual capture histories of

N. vittata, N. pygmaea and N. balstoni tagged in Refuge 1; however, due to the very low numbers of recaptures of the latter two species in Refuge 2, it was only possible to model 121

Figure 5.4. Body locations of visible implant elastomer (VIE) tags. Each fish was

given only two tags using a combination of five different tag colours. Only

the left side of the fish is shown, but the same locations were also used on

the right side of the body giving a total of six potential tagging locations

on the body.

122 the N. vittata population in this pool. As only a subsample of the captured N. vittata was tagged on each sampling occasion, a cumulative set (i.e. treated as batch-marked) of

Lincoln-Petersen population estimates (with Chapman’s modification) for this species was calculated for each pool using data from the multiple mark-recapture periods (see

Lockwood & Schneider 2000).

A number of different models were tested for modelling the population size of N. pygmaea and N. balstoni including a range of time variable (t) and fixed probability of capture and re-captures. However, the heterogeneity time response model of Chao et al.

(1992) (Mth) was selected as it assumed variable probability of capture among individuals as well as sampling periods (to account for habitat and behavioural changes over the sampling period such as declining water levels, changes in foraging and movement behaviour, etc.). Population models were fitted using CAPTURE in the MARK program.

5.3 Results

5.3.1 Distribution of N. pygmaea

In total, 10 of the 43 sites sampled for fish in the Hay River catchment housed N. pygmaea (Figure 5.5A). This result confirmed that the species has a baseflow distribution that is restricted to just a few pools in the mainstem of the Hay River near the Mitchell

River confluence. In addition, the lowermost ~2 rkm of the Mitchell River are utilised during the annual winter/spring flow period. Two other nearby tributaries (i.e. Sheepwash

Creek and Blue Gum Creek) were sampled on several occasions throughout the study period but N. pygmaea was not found to inhabit these streams (Figure 5.5A).

123

A)

B)

Figure 5.5. A) Location of sites sampled for fish in the Hay, Denmark, Kent and Bow

rivers, Western Australia, during the current study (2009-2014), and; B)

map of the Extent of Occurrence (EOO) of Nannoperca pygmaea showing

the outlying population recorded in Lake Smith.

124

A further 50 sites were sampled in other nearby river systems, with N. pygmaea found at 10 sites in the Denmark and Kent rivers; however, the species was not found in the Bow River (Figure 5.5A). In addition, opportunistic sampling uncovered a fourth sub- population of N. pygmaea at Lake Smith, ~200 km west of the nearest known population in the Kent River (Figure 5.5B). The current known EOO for the species is 3,420 km2 and the AOO is 10 km2. Nannoperca pygmaea represented just 2.2% of the total number of native fish captured during the study. It comprised ~18%, ~8% and ~4% of the total number of native fishes caught in Lake Smith, the Denmark River and the Kent River, respectively, cf. only 1.5 % in the Hay River system.

5.3.2 Biological characteristics of N. pygmaea

Nannoperca pygmaea was reproductively active during winter-spring in the Hay/Mitchell system with running ripe individuals captured in August and September over two consecutive years (Figure 5.6). The highest prevalence of spawning individuals was recorded in August of both years (Figure 5.6). It had a much shorter breeding period than both N. balstoni and N. vittata in 2013, with the breeding periods for the three species staggered across the winter-spring period (Figure 5.6). Nannatherina balstoni bred earliest, commencing in May prior to the onset of winter flows, with most individuals spent by September, while N. pygmaea commenced breeding in August and was spent by

September. Nannoperca vittata had a more protracted breeding period that extended from

August to November. Unfortunately, a lack of sampling during the early part of the breeding period (i.e. May-July) precluded making the same comparison in 2014; however, the staggered breeding trend was also evident in that year, with each species reproductively spent in approximately the same month as for 2013 (Figure 5.6). 125

2013

NB 2013 NP NV 100 800

600 80 =12 n

60 400 =25

n =2

40 n 200 Discharge (ML/day)

% in breeding conditionbreeding in% 20

0

=22 =380 =22 =302 n =602 =9 n =24 n =513 n n n n 0 NS n M J J A S O N u u u e o a n l c y g p t v 2014 100 800

2014 80 600

60 400

=4 =13 n 40 n 200

Discharge (ML/day)

% in condition breeding % 20

=158 0 =32 n =69 n =68 =216 n =20 n n n 0 NS NS M J J A S O N u u u e o a n l c y g p t v

Figure 5.6. Percentage of N. balstoni (NB), N. pygmaea (NP), and N. vittata (NV)

sampled from the Hay/Mitchell system during 2013 (above) and 2014

(below) that were in breeding condition (i.e. ‘running ripe‘). Numbers of

each species collected in each month are given. Daily discharge in the

Mitchell River is plotted as a line graph. NS - fish not sampled during the

indicated month.

126

Histological examination of the gonads of two females in breeding condition revealed the presence of oocytes at different cytological stages (Figure 5.7), indicating that N. pygmaea is likely to be a serial spawner (i.e. produces numerous batches of eggs during the breeding season). While early stage chromatin nucleolar and perinucleolar stage oocytes dominated numerically, oocytes at later developmental stages were also present in relatively large numbers. Yolk granule oocytes (~150-400 µm in diameter) were almost as numerous as those at earlier developmental stages, whilst ~2.4 and ~4.2% of oocytes present comprised hydrating oocytes (~500-800 µm in diameter) and fully hydrated ova (~800-900 µm in diameter), respectively. One of the ovaries examined histologically also contained post ovulatory follicles, indicating that ova had been released by the female prior to sampling (Figure 5.7).

Counts of large (i.e. >700 µm diameter) oocytes in the ovaries of the five spawning females examined ranged between 27 and 105 (i.e. mean batch fecundity 60.6

± 13.2 S.E.). Numbers of oocytes in the smaller-size categories estimated by extrapolating from sub-sample counts were as follows: 500-700 µm diameter, 60-168

(mean = 61.0 ± 12.9 S.E.); 350-500 µm diameter, 116-304 (mean = 246.6 ± 33.4 S.E.);

200-350 µm diameter, 246-532 (mean = 362.6 ± 51.8 S.E.); <200 µm diameter, 1134-

2494 (mean = 1810.8 ± 231.8 S.E.). The largest individual female N. pygmaea examined for fecundity (51 mm TL) appeared capable of releasing up to ~3,100 eggs during a single breeding season. Note, however that this fish was substantially smaller than the largest individual captured during the study (66 mm TL), therefore maximum batch and total fecundity for the species is likely to be higher.

127 Post-ovulatory follicle Hydrating

A) Chromatin nucleolar

Perinucleolar

Late yolk vesicle

Early yolk vesicle

B)

Figure 5.7. A) Histological section and B) dissection of the ovary of a spawning

(‘running ripe’) Nannoperca pygmaea captured from the Hay River system

in September 2013, showing the various stages of oocyte development.

128

In total, 41 individual N. pygmaea were captured in spawning condition during the study. Of these, 18 were female (39-56 mm TL, modal length = 43 mm TL) and 23 were male (39-51 mm TL, bimodal lengths of 43 and 47 mm TL) (Figure 5.8). The majority of spawning N. pygmaea were captured in the Mitchell River, but five individuals were also captured in the Denmark River.

The otoliths of 10 female N. pygmaea that ranged in size between 41 and 49 mm

TL had two hyaline zones indicating that fish were ~2 years old (Figure 5.9A-C). Narrow hyaline bands were apparent within the opaque zones of most of the otoliths from these 2- year old fish, suggesting that they had undergone brief growth checks during the period of sustained faster growth in each year (Figure 5.9A-C). It was unclear whether the largest specimen examined (51 mm TL) had three or four hyaline zones; the fourth may have been a significant growth check (Figure 5.9D). This configuration of annuli suggests that the fish was at least 3-years old or perhaps in its 4th year.

The maximum size attained by N. pygmaea differed among the three main sub- populations (Figure 5.10). The largest individuals recorded in each catchment measured

51 mm TL (Kent), 62 mm TL (Hay), and 66 mm TL (Denmark). A cohort in the Hay

River was clearly evident in July 2013 (indicated by blue arrows on Figure 5.10), comprising fish that were likely in their second year (i.e. 1+, see results of otolith examination above for rationale). This cohort attained a modal length of ~43 mm TL in

August 2013 (Figure 5.10), as the end of the second year of life approached; assuming an early-September hatching (i.e. approximately the mid-point of the spawning period; see

Figure 5.6). Thus, as the lengths of fish in this cohort fell mostly within the size range of individuals in spawning condition (see Figure 5.8), it is likely that N. pygmaea in the Hay

River does not reach sexual maturity until it approaches two years of age. 129

6

Males

4

Frequency 2

0 36 38 40 42 44 46 48 50 52 54 56 58 Total length (mm)

6 Females

4

Frequency 2

0 36 38 40 42 44 46 48 50 52 54 56 58 Total length (mm)

Figure 5.8. Length-frequency histograms of male (upper) and female (lower)

Nannoperca pygmaea that were captured in spawning condition, pooled

for all sites sampled across the entire study period.

130

Figure 5.9. Sagittal otoliths of 2-year old (A, B, C) and 3-year old (D) Nannoperca

pygmaea viewed under reflected light. All specimens were females in

breeding condition (‘running ripe’).

131 2013 2014

15 30 March March 15 10 Denmark Not sampled 10 Kent Frequency Frequency 5 Hay 5

0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70

15 15 April April 10 10 Frequency Frequency 5 5

0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70

15 35 May May 20 10 15 10 Frequency Frequency 5 5 0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70

15 15 July July 10 10 Not sampled

Frequency 5 Frequency 5

0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70

15 15 August August 10 10 Frequency Frequency 5 5

0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70

15 15 September September 10 10 Frequency Frequency 5 5

0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70

15 15 October October

10 10 Frequency Frequency 5 5

0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70

15 15 November November

10 10 Frequency Frequency 5 5

0 0

10 15 20 25 30 35 40 45 50 55 60 65 70 10 15 20 25 30 35 40 45 50 55 60 65 70 Total length (mm) Total length (mm)

Figure 5.10. Monthly length-frequency histograms for Nannoperca pygmaea in the

three main sub-populations. Blue arrows indicate a cohort of fish likely to

have hatched in Spring 2011.

132

By March 2014, this cohort comprised the majority of the population sampled in the Hay River, and a cohort of young-of-year fish (modal length ~30 mm TL) was also present (Figure 5.10). Note that the above average numbers of fish captured during this month was due to the intensive sampling involved with the mark-recapture component of the project (see below). Unfortunately, a lack of sampling in the Hay River from May to

July 2014 made it impossible to continue tracking these cohorts, but the absence of larger fish (i.e. ≥55 mm TL) after September 2014 sampling (Figure 5.10) suggests that, at least in the Hay River sub-population, a substantial proportion of adult N. pygmaea may perish during or shortly after the annual breeding period.

Although sampling was only conducted sporadically in other river systems, the

Kent sub-population had similar population demographics to that in the Hay. Two distinguishable cohorts were present in the Kent River in May 2014, but the older cohort had mostly disappeared from samples by October/November (Figure 5.10). Individuals up to 65 mm in length were present in the Denmark River sub-population (Figure 5.10).

Examination of the contents of the intestinal tracts of 12 adult females revealed that a range of macroinvertebrates including terrestrial larvae (i.e. dipterans), copepods and ostracods had been consumed. These fish were captured in either August or

September 2013 from the Mitchell River.

5.3.3 Migrations of pygmy perches

In the Hay River system, N. pygmaea was repeatedly sampled from the intermittent

Mitchell River during the annual flow period over a number of years, but was found to be absent from all permanent refuge pools in this tributary during baseflow. Despite 133 extensive sampling effort during baseflow periods over two years throughout the Hay catchment, the species was only ever captured in four refuge pools in the Hay River mainstem that were all located within 400 m of the Mitchell River confluence.

A comparison of fyke net catch data between sites in the Hay River mainstem and the Mitchell River revealed that upon commencement of winter flow in July 2013, N. pygmaea were predominantly captured in the Hay River (Figure 5.11), although > 50% of these fish were captured at the Hay-Mitchell confluence. As mean daily discharge increased, however, the situation reversed, with N. pygmaea captures made exclusively in the lower Mitchell River in August 2013 (Figure 5.11). Mitchell River captures continued to outnumber those in the Hay through to October, but in November, as seasonal flows abated, the majority of N. pygmaea were captured in the Hay River (Figure 5.11).

Unfortunately, a comparison could not be made with data from the early part of the flow period in 2009 (a year of high discharge) as sampling did not take place until August, at which time N. pygmaea were exclusively captured in the Mitchell River (Figure 5.11).

However, a similar trend to that observed in the 2013 data was apparent at the end of spring flows in 2009, and again in 2014, although numbers captured in the Hay exceeded those in the Mitchell a month earlier (i.e. October) than they did in 2009 and 2013, most likely due to the earlier cessation of streamflow resulting from the significantly lower rainfall and surface discharge in 2014 (Figure 5.11). In the Hay sub-population, approximately 97% of the total number of N. pygmaea that were found to be in breeding condition were captured from the Mitchell River.

Similar trends were evident for N. balstoni, with more captures generally made at

Mitchell River sites during months of higher discharge, whilst captures in the Hay River became more prevalent in November as flow abated (Figure 5.12). Throughout the

134

Nannoperca pygmaea

0.8 1000 Hay River Mitchell River 800

0.6

600

0.4 400

200

Average number / net hr Average 0.2 Discharge (average ML / day) ML (average Discharge 0

NS NS NS 0.0 9 9 3 3 4 4 09 0 13 13 13 1 1 1 14 14 009 0 00 0 0 0 0 0 0 0 0 2 2 2013 2 2 20 2014 2 p v 2 n y g t c n p ly 2 e o u ly 2 e Ju Aug 2 S Oct 2009 N Dec 2009 Ju Jul A Sep 2 Oc Nov 2013 De Ju Ju Aug 2014 S Oct 2 Nov 201

2009 2013 2014

Figure 5.11. Fyke net captures (average number of individuals.net-1.h-1) of Nannoperca

pygmaea in the Hay and Mitchell rivers during the major flow period in

2009, 2013 and 2014. Average daily surface discharge in the Hay River is

plotted as a line graph. NS – indicates no sampling undertaken during a

particular month.

135

Nannatherina balstoni

0.6 1000 Hay River Mitchell River 0.5 800

0.4 600

0.3 400

0.2 200 Average number / net / hr / net / number Average Discharge (average ML / day) ML (average Discharge 0.1 0

NS NS NS 0.0 9 9 3 3 3 3 3 3 4 4 4 09 0 1 1 1 1 1 14 00 0 01 01 0 01 2 20 20 2014 2 2 2 g 2 p n 2 c 20 n g 2014 p ov 2009 ly e u ct 201 ov July Au Se Oct 2009 N Dec 2009 Ju Ju Aug 20 Sep 20 Oct 20 Nov 2013 D Ju July A Se O N

2009 2013 2014

Figure 5.12. Fyke net captures (average number of individuals.net-1.h-1) of

Nannatherina balstoni in the Hay and Mitchell rivers during the major

flow period in 2009, 2013 and 2014. Average daily surface discharge in

the Hay River is plotted as a line graph. NS – indicates no sampling

undertaken during a particular month.

136 duration of the study, the number of N. balstoni that were captured in spawning condition in the Mitchell River greatly outnumbered those caught in the Hay River (23 vs 7); and females in breeding condition (n = 6) were exclusively captured in the Mitchell River.

The data for N. vittata contrasted that of the other pygmy perch species, with fish captured in the Hay River outnumbering those captured in the Mitchell throughout 2013 and 2014 (Figure 5.13). However in 2009, a year that featured substantially higher stream discharge, the opposite occurred, with more fish captured in the Mitchell River (Figure

5.13). Comparable numbers of breeding females were captured in both the Hay (n = 71) and Mitchell (n =87) rivers over the course of the study.

Eight individual fish (5 × N. pygmaea, 2 × N. vittata, 1 × N. balstoni) that had been captured from refuge pools in the Hay River mainstem during baseflow in March

2014 and marked with VIE tags were later recaptured in the intermittent lower reaches of the Mitchell River at a time of year (August and September) that coincided with the breeding periods of all three species (see Figure 5.6).

5.3.4 Environmental variables

Rainfall was well above average in the study area in 2009 (Figure 5.14), which explains the high stream discharge in the Denmark and Hay rivers that year (Figure 5.15). In 2013, total annual rainfall was close to the long term average, but it fell in an unusual seasonal pattern, with above average rain in autumn, followed by below average rain in early- to mid-winter, and then above average rain in late-winter and into early-spring (Figure

5.14). In 2014, rainfall was around the long term average or slightly less in most

137

60 Nannoperca vittata 1000 Hay River Mitchell River 40 800

20 600

400 6

200

Average number / net / hr net numberAverage / 3 Discharge (average ML / day)Discharge (averageML / 0

NS NS NS 0 9 9 3 3 3 3 3 4 4 4 0 0 1 1 1 1 1 1 1 0 0 0 0 0 0 014 0 0 0 2 2 2 2 2 2 c 2 p v c 2 g p v ly 2009 e ly e o e u e ct 2 Ju Aug 2009 Sep 2009 Oct 2 Nov 2009 D Jun 2013 Ju Aug 201 S Oct 2013 N D Jun 2014 July 2014 A S O No

2009 2013 2014

Figure 5.13. Fyke net captures (average number of individuals.net-1.h-1) of Nannoperca

vittata in the Hay and Mitchell rivers during the major flow period in

2009, 2013 and 2014. Average daily surface discharge in the Hay River is

plotted as a line graph. NS – indicates no sampling undertaken during a

particular month.

138

300 Denmark Rainfall during 250 study period Mean rainfall (1898-2014) 200

150

Rainfall (mm) 100

50

0

J F M A M J J A S O N D J F M A M J J A S O N D J F M A M J J A S O N D 2009 2013 2014

300 Denbarker Rainfall during 250 study period Mean rainfall (1967-2014) 200

150

Rainfall (mm) 100

50

0

J F M A M J J A S O N D J F M A M J J A S O N D J F M A M J J A S O N D 2009 2013 2014

Figure 5.14. Monthly rainfall data during the sampling period at Denmark (upper) and Denbarker (lower), two locations situated centrally in

the study area. Data were sourced from the Bureau of Meteorology (BoM 2015).

Discharge Denmark River 2500 Hay River ) -1 2000

1500

1000

500 Total(ML.d daily discharge

0

J F M A M J J A S O N D J F M A M J J A S O N D J F M A M J J A S O N D 2009 2013 2014

Conductivity 12000 ) -1 10000

8000

6000

4000

2000 Mean daily conductivity(µS.cm 0

J F M A M J J A S O N D J F M A M J J A S O N D J F M A M J J A S O N D 2009 2013 2014

Figure 5.15. Total daily discharge and mean daily conductivity recorded in the Hay (Sunny Glen) and Denmark (Mt. Lindesay) rivers during

the study period. Data were sourced from the Department of Water, Water Information Reporting service (DoW 2015). 139

140 months, but was well below average in August, which is the third wettest month of the year on average (Figure 5.14). The lack of rainfall in August accounts for the low river discharge across the study area in 2014 (Figure 5.15). Discharge was ~2.5 times greater between August and October in 2013 cf. 2014 in the Hay and Denmark rivers (Figure

5.15). Much wetter conditions were recorded in 2009 with considerably higher discharge than both 2013 and 2014 (Figure 5.15). Despite this, very little difference was observed in the CPUE data for N. pygmaea in the Hay-Mitchell system among these three years.

During the study period, conductivity in the Hay River followed a consistent annual pattern of relative stability (6,000-8,000 µS.cm-1) during the baseflow period

(January-May), before increasing sharply to around 12,000 µS.cm-1 with the onset of the first major flows of the year (Figure 5.15). Conductivity declined steadily over winter, reaching a minimum of ~3,000-4,000 µS.cm-1 in late spring, before gradually rising with the onset of summer (Figure 5.15). The conductivity trend was similar in the Denmark

River, but much less pronounced in magnitude (Figure 5.15). Data for the Kent River were not available from the Department of Water during the study period, but similar conductivity levels to those in the Hay River mainstem were recorded during the study using a handheld water quality meter (see Figure 5.16).

5.3.5 Population size estimation in Hay River baseflow refuges

Totals of 66, 518 and 14 N. pygmaea, N. vittata, and N. balstoni, respectively, were marked in two refuge pools in the Hay River during the five sampling events between

March and April 2014 (Tables 5.1-5.5). Nannoperca vittata dominated both refuge pools numerically; the population of this species was estimated at 8,117 (lower/upper 95% CI

2,289-13,944) and 6,423 fish (lower/upper 95% CI 4,863-7,984) in Refuge Pool 1 and 2, 141

Baseflow 25 Water temperature 100 Dissolved oxygen (%) Peakflow Shoulder

20 80

15 60

10 40 Temperature (°C) Dissolved (%) oxygen

5 20 DRY IN BASEFLOW 0 0 DRYBASEFLOW IN

Hay Mitchell Denmark Kent Hay Mitchell Denmark Kent

8 Salinity 10000 Conductivity

8000 6 ) -1

6000

4 ppt

4000 Conductivity (µS.cm 2 2000 DRY IN BASEFLOW 0 DRY IN BASEFLOW 0

Hay Mitchell Denmark Kent Hay Mitchell Denmark Kent

10 pH 10000 Total dissolved solids

8 8000

6 6000 pH

4 TDS (ppm) 4000

2 2000 DRY IN BASEFLOW 0 0 DRY IN BASEFLOW

Hay Mitchell Denmark Kent Hay Mitchell Denmark Kent

Figure 5.16. Mean (± 1 S.E.) environmental variables measured during 2014 in

baseflow (March-April), peakflow (August-September) and shoulder flow

(November) conditions in sections of the Hay, Mitchell, Denmark and

Kent rivers, Western Australia, occupied by Nannoperca pygmaea.

142

Table 5.1. Number of Nannoperca vittata captured, marked, and not marked in Refuge Pool 1

Sampling occasion Days total mark total Total Total not marked Total not marked Total +recap recaptures marked or recap + recaps captured 6/03/2014 0 30 0 30 1,048 1,048 1,078 13/03/2014 7 41 3 38 604 607 645 26/03/2014 13 40 7 33 1,462 1,469 1,502 9/04/2014 14 37 7 30 200 207 237 16/04/2014 7 5 5 0 363 368 368

Table 5.2. Number of Nannoperca vittata captured, marked, and not marked in Refuge Pool 2

Sampling occasion Days total mark total Total Total not marked Total not marked Total +recap recaptures marked or recap + recaps captured 20/02/2014 0 120 0 120 1,453 1,453 1,573 6/03/2014 17 90 17 73 784 801 874 13/03/2014 7 72 12 60 207 219 279 26/03/2014 13 87 19 68 515 534 602 9/04/2014 14 72 6 66 70 76 142 16/04/2014 7 51 51 0 809 860 860

Table 5.3. Number of Nannoperca pygmaea captured, marked, and not marked in Refuge Pool 1

Sampling occasion Days total mark total Total Total not marked Total not marked Total +recap recaptures marked or recap + recaps captured 6/03/2014 0 17 0 17 1 1 18 13/03/2014 7 14 1 13 0 1 14 26/03/2014 13 29 10 19 2 12 31 9/04/2014 14 7 3 4 0 3 7 16/04/2014 7 5 5 0 7 12 12

Table 5.4. Number of Nannoperca pygmaea captured, marked, and not marked in Refuge Pool 2

Sampling occasion Days total mark total Total Total not marked Total not marked Total +recap recaptures marked or recap + recaps captured 20/02/2014 0 3 0 3 0 0 3 6/03/2014 17 1 0 1 0 0 1 13/03/2014 7 1 0 1 0 0 1 26/03/2014 13 7 0 7 0 0 7 9/04/2014 14 1 0 1 0 0 1 16/04/2014 7 1 1 0 10 11 11

143

144

Table 5.5. Number of Nannoperca balstoni captured, marked, and not marked in Refuge Pool 1

Sampling occasion Days total mark total Total Total not marked Total not marked Total +recap recaptures marked or recap + recaps captured 6/03/2014 0 7 0 7 0 0 7 13/03/2014 7 5 2 3 0 2 5 26/03/2014 13 7 4 3 1 4 8 9/04/2014 14 0 1 0 1 1 1 16/04/2014 7 1 1 0 1 2 2

145

respectively. The Mth model estimated the number of N. pygmaea and N. balstoni in

Refuge Pool 1 at 90 (±15.5 SE) and 26 (±11.7 SE), respectively. Numbers of N. pygmaea and N. balstoni captured and marked in Refuge Pool 2 were insufficient to generate population estimates.

5.4 Discussion

5.4.1 Distribution of N. pygmaea

The hypothesis that N. pygmaea would have a more widespread distribution than previously known was confirmed during the study, as specimens were captured at five sites in the Denmark River, which shares a watershed boundary with the Hay/Mitchell system (Figure 5.5A). Initially, it was unclear whether these fish were in fact N. pygmaea, as they differed slightly in appearance and attained a larger maximum size than fish previously captured from the Hay River; however, allozyme analysis demonstrated that the two sub-populations belonged to the same phylogenetic lineage (M. Adams, unpublished data). Additional sub-populations were also found during the study in the

Kent River, which neighbours the Denmark system, and in Lake Smith, part of the

Donnelly River catchment. Allozyme testing has not yet been performed on these sub- populations; however, both share a morphological resemblance to N. pygmaea from the

Hay River.

Under IUCN Red List assessment criteria, N. pygmaea qualified as Endangered due to its limited EOO (i.e. <5000km2) and AOO (i.e. <500km2). The low number (i.e. ≤

5) of known sub-populations and the fragmentation of the distributional range provided further justification for this categorisation. The core of the species distribution centres

146 around three adjacent catchments (i.e. Hay, Denmark, Kent) with an outlying record from

Lake Smith. The latter discovery was unexpected given that this site is located ~200 km from the nearest known population in the Kent River and is a lacustrine habitat, whereas other populations are known only from lotic systems. Many of the lacustrine and riverine habitats between Lake Smith and the Kent River have not been surveyed recently for fish; therefore it is possible that N. pygmaea is more widespread than the results of this study indicate. Note, however, that another, as yet, undescribed species of Nannoperca inhabits drainages in the vicinity of Broke Inlet (J. Murphy et al., unpublished data) and surveys of four sites in the Bow River, a drainage system situated in the notable distribution

“gap” between Kent River and Lake Smith, failed to detect N. pygmaea. Regardless, further targeted surveys between Kent River and Lake Smith, particularly of freshwater lacustrine habitats, are recommended in order to elucidate the EEO and AOO of this rare species with greater precision. Emerging technologies such as environmental DNA

(Thomsen et al. 2012) may have a role to play in future survey work aimed at elucidating the full extent of the distribution of N. pygmaea and other threatened fishes in the region.

5.4.2 Biology and migration

Despite the restriction on the number of fish that could be destructively sampled, guided by Animal Ethics regulations, valuable data on several key biological traits were gathered during the study. This included the timing and strategy of reproduction, fecundity, age and growth, diet, and seasonal migration patterns. As hypothesised, N. pygmaea shared a number of biological traits in common with its sympatric congener N. vittata (see Pen &

Potter 1991a), but some notable differences were also revealed.

147

Prey items consumed by the 12 individuals examined during this study of N. pygmaea closely matched dietary preferences previously determined for N. vittata (Pen &

Potter 1991a) and juvenile N. balstoni (Morgan et al. 1995), although adults of the latter species switch to a specialised diet consisting almost exclusively of terrestrial fauna such as and spiders that alight on the surface of the water. Ontogenetic and seasonal patterns of dietary composition remain unknown in N. pygmaea and merit further investigation.

Examination of otoliths of N. pygmaea collected from the Hay River in

August/September 2013 revealed that those sized 41-49 mm TL were 2 years old, and a single fish of 51 mm TL was probably 3 years old. Thus, it appears that the growth rate of this population is slower than that recorded by Pen & Potter (1991a) for N. vittata, which reaches 40-45 mm TL at 1 year of age and 51-53 mm TL at 2 years of age (Pen & Potter

1991a), but similar to that recorded for the eastern Australian species, Nannoperca australis, which attains a length of ~35 mm TL at the end of its first year (Llewellyn

1974). Contrastingly, N. balstoni grows considerably faster, attaining 60-70 mm TL at age 1 (Morgan et al. 1995).

The longevity of N. pygmaea was not determined in this study but the presence of specimens ~65 mm TL in the Denmark River suggests a maximum age well in excess of

3 years, noting that N. vittata between 60 and 70 mm TL were 5 years old in the Collie

River (Pen & Potter 1991a). The prevalence of larger individuals in the Denmark River sub-population cf. other sub-populations may be attributable to the lower salinity levels of baseflow refuge habitats in that catchment, but this hypothesis was not tested. Studies have shown that increasing salinity inhibits growth rates in a number of freshwater fish species in other parts of the world (e.g. Likongwe et al. 1996; Wang et al. 1997; Küçük

2013), and this phenomenon could be influencing growth and mortality rates of N.

148 pygmaea in the secondarily salinised Hay and Kent rivers. Moreover, dissolved oxygen levels during baseflow in the latter systems, were lower than those recorded in the

Denmark River. This may have been another factor contributing to differences in population demographics among these systems as growth rates of some fish species (e.g.

Ictalurus punctatus) have been shown to decrease with exposure to low dissolved oxygen concentrations (Andrews et al. 1973). In order to definitively ascertain if fish live longer and/or grow faster in the Denmark system, additional destructive sampling would be required in order to undertake otolith analysis so as to determine age at length of the largest specimens in each of the sub-populations. Controlled laboratory experiments would also contribute valuable data to test hypotheses surrounding the potential influence of environmental variables such as salinity and dissolved oxygen concentration on growth rates and physiological tolerance thresholds of N. pygmaea.

Determination of age and growth characteristics of N. pygmaea proved difficult due to inconsistent numbers of fish in each size class (cohort) sampled from month to month. Whilst this was at least partly attributable to the extreme rarity of the species, it was also potentially an artefact of differences in migrational behaviour of the different age classes in the Hay/Mitchell system. Sampling during the annual flow period (i.e.

~June to November) in 2013 and 2014, resulted in captures of N. pygmaea exclusively in the Mitchell River, and of these, only ~2% of fish belonged to what is presumed to be the young-of-year (i.e. <35 mm TL) age class. Therefore it appears that 0+ fish do not typically migrate laterally into the Mitchell River. The lack of captures in the Hay River mainstem during the annual flow period was probably a reflection of the extremely low relative abundance of the species, compounded by the difficulties associated with sampling fish at high water levels in mainstem habitats.

149

Nannoperca pygmaea <39 mm TL were not found to be in spawning condition during the study, therefore it appears that sexual maturity is not attained until fish are in their second year. This contrasts with N. vittata, N. balstoni and other freshwater fish species in the region that reach maturity at the end of their first year (e.g. Pen & Potter

1991a, b; Pen et al. 1991; Morgan et al. 1995, 1998). It may also help to explain why the species has a much lower relative abundance than N. vittata, as fish have a longer period of exposure to potential mortality (e.g. predation, disease, physiological stress, competition) before reproducing.

The timing of reproduction was partitioned amongst the three sympatric pygmy perch species in the Hay River system across the winter/spring period. Temperature appears to be a vital breeding cue for N. balstoni, as its peak breeding period (July) roughly coincided with the annual minimum water temperature recorded in the Hay River catchment. Morgan et al. (1995) similarly reported an unusually early breeding period for

N. balstoni in lentic pool habitats located in the acidic peat flats near Windy Harbour. The breeding period of N. vittata was protracted, extending from August to November with a peak in October (mid-spring) that coincided with late season flow, rising water temperatures (> 12-13° C) and increasing day length. Peak spawning activity was also found to occur from late-September to October in the population of N. vittata (Pen & Potter 1991a), and is similarly timed for the eastern Australian species N. australis (Llewelyn 1974). Spawning activity in N. pygmaea commenced and peaked during the month of August, coinciding with the first significant flow pulses of the year, and was finished by October. Pen and Potter (1991a) hypothesised that, inter alia, differences in spawning times reduces the potential for interspecific competition amongst freshwater fishes. The observed temporal partitioning of peak spawning activity among the three pygmy perch species in the current study potentially adds further evidence to

150 support this hypothesis; however to confirm this, more research is needed on the habitat requirements and dietary preferences of these species, particularly the early life history stages.

The prevalence of potamodromy varied among species, with strong evidence gathered that lateral migrations from main channel refuges into seasonally flowing tributaries are vital for reproduction in the Hay River populations of both N. pygmaea and

N. balstoni. During the annual flow period, 100% of running ripe females and over 95% of total captures for these two species were made at sites in the ephemeral Mitchell River, suggesting that potamodromy is obligate for populations of both species in the Hay River system. Moreover, individuals of all three pygmy perch species that had been captured and tagged in a Hay River refuge pool in summer were recaptured in the lower Mitchell

River during the annual flow period, with all but one of these fish either running ripe or recently spent. This is the first time that potamodromous migration of freshwater fishes has been definitively proven using mark-recapture methods in south-western Australia.

Further evidence of potamodromy for N. pygmaea was gathered in the Denmark and Kent catchments. In the Denmark River, fish were sampled from three permanent refuge sites in a headwater tributary (‘Little Pygmy Creek’; see Figure 5.2). Aerial surveying during baseflow demonstrated that permanent refuge habitat was absent from the Denmark River mainstem near the confluence with this tributary, yet a number of running ripe individuals were captured in the Denmark mainstem ~1.2 km downstream of the tributary confluence during the study. Therefore, some fish probably migrated downstream from refuges in ‘Little Pygmy Creek’ to spawn at sites in the Denmark River mainstem, i.e. the opposite direction to that observed in the Hay River sub-population.

151

In the Kent River during baseflow 2014, N. pygmaea was captured in high relative abundance (cf. other sub-populations); however, it was absent from the same site during the winter breeding period. It was hypothesised that these fish had undertaken a spawning migration, but unfortunately the destination of migration could not be located in the Kent

River to test the hypothesis. Interestingly, in November, as seasonal flow abated, N. pygmaea (~45% of which were young-of-year) re-appeared at the same baseflow refuge site where it had not been captured in winter, suggesting that a potamodromous breeding migration followed by a retreat to permanent refuge habitat had occurred.

In contrast to N. pygmaea, potamodromy appeared to be facultative rather than obligatory for the Hay River population of N. vittata. Whilst the ratio of running ripe female N. vittata captured at sites in the Hay and Mitchell varied inter-annually, the overall numbers captured in these two rivers during the study were similar (71 in the Hay vs 87 in the Mitchell). This was consistent with the findings of Pen & Potter (1991a), who documented a similar situation in the Collie River with a proportion of the N. vittata population either moving into floodwaters adjacent to mainstem refuge habitats or into seasonally flowing lower order tributaries during the annual breeding period.

Importantly, neither N. pygmaea nor N. balstoni were found to utilise refuge pools in the Mitchell River, instead, retreating to Hay River mainstem refuges despite the fact they were slightly salinised; noting that elevated salinities are known to be deleterious to

N. balstoni (Beatty et al. 2011). The exclusive use of the non-salinised, seasonally flowing Mitchell River for spawning by N. pygmaea and N. balstoni in the Hay system suggests an inability of the larval and post-larval stages of these species to survive in salinised habitats, unlike N. vittata, which apparently has a higher salinity tolerance during its early life history. This is not surprising given that in laboratory trials of acute salinity tolerance in adult fish, N. vittata was found to tolerate salt concentrations almost

152 twice that of N. balstoni (Beatty et al. 2011). The salinity tolerance of N. pygmaea is unknown but may parallel that of N. balstoni judging from its potamodromous migration pattern in the Hay River system.

Histological analysis of mature female gonads proved that N. pygmaea is a serial spawner (i.e. females produce multiple batches of eggs in a breeding season). This is a reproductive trait shared with N. vittata (Shipway 1949; Pen & Potter 1991a) and is a common strategy within the Australian freshwater fish fauna (see King et al. 2013 for examples). This mode of reproduction is an effective means by which small-bodied fishes are able to maximise their total egg production during a breeding season (Prince & Potter

1983; Molsher et al. 1994).

5.4.3 Population estimates

The VIE mark-recapture program conducted during this study represented the first quantification of the abundance of freshwater fishes in natural refuge pools in south- western Australia. Nannoperca vittata greatly outnumbered both N. pygmaea and N. balstoni, highlighting the precarious conservation status of the latter species in the Hay

River system. The mean batch fecundity for the five N. pygmaea females examined was considerably less than that recorded by Pen & Potter (1991a) for similarly aged N. vittata

(~60 vs ~370), which may have partly accounted for the observed disparity in population size between these two species. The later age of sexual maturity in N. pygmaea may have also contributed to this result; however, this does not explain why N. balstoni is so rare, as it attains maturity in its first year (Morgan et al. 1995). The similarly small populations of N. pygmaea and N. balstoni suggest that the requirement for these species to breed in non-salinised habitats may be a major factor in limiting their population sizes in the Hay

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River. Evidently, the habitat requirements for spawning and early development of N. vittata are not as specific and the ability of this species to breed in the Hay River mainstem as well as in lower order tributaries such as the Mitchell River, gives this species an advantage over sympatric pygmy perch species in terms of its reproductive and recruitment capacity. This is the most parsimonious explanation for the numerical dominance of N. vittata over both N. pygmaea (~100 fold less abundant) and N. balstoni

(~300 fold less abundant) in the smaller of the two refuge pools.

One of the more interesting results to emerge from the population estimates was the larger population of N. vittata in the smaller (by surface area and volume) of the two refuge pools. This result suggests that pool volume and depth per se are not reliable indicators of carrying capacity for freshwater fishes in south-western Australia. The smaller pool contained a higher proportion of large woody debris, which is known to provide important habitat for pygmy perch species (Merrick & Schmida 1984; Bond &

Lake 2003; Knight & Arthington 2008) and showed less pronounced stratification of temperature and dissolved oxygen concentration, which probably accounted for this result. Observations made while snorkelling in this pool confirmed that fishes were congregating along the margins of the pool and were rarely observed below a depth of

~1.0 m. These results have implications for the barrier prioritisation process (see Chapter

3) as they indicate that habitat “quality” as opposed to “quantity” may be more important in determining the carrying capacity of fishes in refuge habitats in this region.

5.4.4 Management implications and future research requirements

This study has elucidated a number of key life history traits for N. pygmaea and located three previously undetected sub-populations in southern Western Australia. These data

154 were used to formally assess the conservation status of the species under International

Union for the Conservation of Nature (IUCN) criteria, culminating in the recent listing of

N. pygmaea as Endangered (Schedule 2) on the Western Australian Government’s

Wildlife Conservation (Specially Protected Fauna) Notice 2015.

Direct evidence of population and/or range decline is lacking due to its recent discovery; however, the extremely constrained AOO suggests a major historical decline within the catchments that it currently occupies given the lack of significant natural barriers to movement in these catchments. The most parsimonious explanation for this is secondary salinisation, which has also been implicated in similar declines of freshwater fish species (including N. balstoni) in the Blackwood River (Morgan et al. 2003; Beatty et al. 2011, 2014a). The upper catchments of the Kent and Hay rivers have been extensively cleared, which has led to elevated salinity in headwater and mainstem habitats of these systems (Pen & Apace 1995; De Silva et al. 2006; Master 2009; DoW

2015). However, the input of fresh runoff from tributaries arising in extensively forested sections of these catchments has a moderating effect on mainstem salinities, allowing salt sensitive species such as N. pygmaea and N. balstoni to persist within narrow spatial ranges. For instance, the restriction of these species in the Hay system to the vicinity of the Mitchell River confluence emphasises how vitally important this tributary is in sustaining these populations. An analogous situation exists in the Blackwood River, where a small number of non-salinised tributaries in the forested middle reaches provide refuge for stenohaline fishes (e.g. N. balstoni and G. munda) within a secondarily salinised system that is dominated by halo-tolerant species including alien Gambusia holbrooki and the estuarine vagrant Leptatherina wallacei (Morgan et al. 2003).

Evidence gathered during this study also revealed a similar situation to exist in the secondarily salinised Denmark River, with N. pygmaea and N. balstoni restricted to a

155 single tributary stream that arises in forested section of the catchment. A major point of difference, however, between the Denmark and other salt-affected rivers such as the Hay and Kent is the considerable management effort that has gone into reversing the rising salinity trend. A remediation strategy ‘to protect and restore key water resources to ensure salinity is at a level that permits safe, potable water supplies in perpetuity’

(Government of Western Australia 2000) was implemented in the Denmark catchment in the late-1980s involving tight restrictions on land clearing and extensive reforestation of cleared farmland (Ward et al. 2011). Salinity levels have been declining since 1990 and this trend is projected to continue (Ward et al. 2011), therefore salt-sensitive fauna such as N. pygmaea and N. balstoni may be expected to expand in range and abundance in this system in the future. The reversal of secondary salinisation in the Denmark River serves as a model for countering this major threat to aquatic biodiversity in south-western

Australia.

Another significant threat to fishes in the region that may not be so readily countered is anthropogenic climate change. Beatty et al. (2014a, b) found the numbers of potamodromous fishes migrating into two lower order tributaries of the Blackwood River to be positively correlated with annual stream discharge. Supporting evidence was not gathered during the current study due to a temporally limited dataset, but these findings are nevertheless concerning given the reliance of N. pygmaea and N. balstoni on potamodromous migration to complete their life cycles, especially in salinised systems.

Projected reductions in rainfall and surface flow are likely to have major implications for reproduction and recruitment of these species (see for e.g. Morrongiello et al. 2011;

Beatty et al. 2014a).

Human exploitation of water resources is also likely to contribute to future stream flow declines across south-western Australia (CSIRO 2009a, b; Morrongiello et al.

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2011). Ironically, it was baseline survey work commissioned for a proposed scheme to transfer water from the Mitchell River into Quickup Dam (part of the Denmark catchment) to augment the Denmark municipal water supply that led to the initial discovery of N. pygmaea (Morgan et al. 2008). This scheme was shelved once the existence of N. pygmaea became known (D. Morgan pers. comm.), but it might be considered again as human water demands inevitably increase in this growing region. The potential impacts on surface discharge in the Mitchell River and the ramifications for potamodromous fishes such as N. pygmaea and N. balstoni (e.g. restricting spatial and temporal access to critical habitats) would likely be significant.

Natural refuge pools are also likely to be seriously impacted in the future due to climate change. Reductions in quantity and/or quality of available refuge habitat has the potential to cause significant population declines, particularly for rare and geographically restricted species such as N. pygmaea. In south-eastern Australia, the recent prolonged

“millennium drought”, resulted in the loss of natural refuges, thus endangering some populations of Mogurnda adspersa, Nannoperca obscura, and N. australis. The situation became so dire that interventionary actions (e.g. captive stock maintenance and breeding programs) were required to prevent the loss of these evolutionary significant units

(Hammer et al. 2013, 2015).

At this stage, similar actions are probably not warranted for N. pygmaea; however, research into captive breeding may be prudent should circumstances change in the future. Translocation of specimens into less vulnerable refuges is another conservation action that has been used for threatened fish species in other parts of

Australia (e.g. Hammer et al. 2013, 2015; Lintermans et al. 2015). The discovery of robust populations of N. pygmaea in two artificial on-stream ponds (used for fire management) in the Denmark River catchment was an important discovery as it

157 demonstrated that the species is capable of thriving in artificial habitats. Numerous other artificial ponds housing native fishes (sans N. pygmaea) were sampled during the study and such sites could potentially be ‘seeded’ with translocated breeding stocks of N. pygmaea as a more cost-effective approach to conservation of the species cf. captive breeding.

The creation of additional strategic artificial habitats is strongly recommended, particularly in the secondarily salinised Hay River catchment, which houses the most spatially restricted sub-population of N. pygmaea. The lower reaches of the Mitchell

River in the vicinity of what is presumed to be the critical spawning/nursery habitat for this population (and that of N. balstoni) would make an ideal candidate site for the construction of such a refuge. Intra-basin translocation of N. pygmaea could also potentially be trialled in a number of existing artificial waterpoints in the Hay catchment, including the pool associated with the upper Mitchell River gauging station weir. This barrier was identified as a high priority for consideration of remediation works to improve fish passage in this study (see Chapter 3). However, the establishment of a translocated population of this endangered species would clearly increase this barrier’s ecological value in providing important artificial refuge habitat, thus necessitating a re- calculation of its prioritisation score and likely affecting its final position in the overall ranking (see Chapters 2 and 3).

Sheepwash Creek, a tributary of the Hay that shares a number of physical and hydrological characteristics with the nearby Mitchell River, offers another promising site for potential intra-basin translocation. This tributary has a forested catchment and is not impacted by secondary salinisation; however, a significant instream barrier (ford) exists within 50 m of its confluence with the Hay River. Unsurprisingly, neither N. pygmaea nor N. balstoni were captured upstream of this barrier during the study as it presumably

158 blocked potamodromous migration of these species into this tributary. Establishment of translocated populations of these species in a newly created artificial refuge above the barrier could potentially facilitate access to significant amounts of additional spawning and nursery habitat within this non-salinised tributary. Modification of the barrier or installation of an appropriately designed fishway could also potentially allow access to and from additional baseflow refuges in the Hay mainstem, thus potentially expanding the AOO of these fishes in the Hay River system.

Further research into artificially created aquatic habitats (e.g. optimal design parameters, micro-habitat requirements, and appropriate spatial distribution across the distributional range) is strongly recommended as a means of offsetting the potential future loss of natural refuge habitat due to climate change. This approach, of course, has the potential to benefit, not only threatened fishes such as N. pygmaea, but also entire aquatic ecosystems (Robson et al. 2011) and perhaps dependent terrestrial fauna as well, although artificial water points can become hotspots of activity for environmental pests including feral cats and foxes (e.g. James et al. 1999), and the risk of this would need to be considered as part of management planning.

The establishment of additional populations of threatened native fishes in artificial habitats would also serve to reduce the susceptibility of these rare species to catastrophic perturbations such as the introduction of an alien aquatic species (e.g. a predator or a disease) or a massive mortality event (fish kill). Fish kills can result from point- (e.g. chemical spills) and non-point source pollution (e.g. nutrient-ladened runoff from agricultural/urban areas) (Cooper 1993; Hodgkin & Hamilton 1993). Whilst the likelihood of the former is low, the risk of a eutrophication induced fish kill does exist in the rivers occupied by N. pygmaea as agriculture is practised in all of these catchments to some extent. Large quantities of nutrient and sediment are discharged annually in both

159 the Hay and Denmark rivers (Pen & Apace 1995). Moreover, stochastic inland rainfall events in secondarily salinised systems can also cause fish kills due to the flushing of high concentrations of salt, especially during summer (e.g. Allen 2002). This is a concern for N. pygmaea as summer rainfall is projected to increase in both frequency and intensity in the future (CSIRO 2009a). Presently, the acute and chronic salinity tolerance levels of all life stages of N. pygmaea are unknown and this knowledge gap should be addressed urgently, especially given the inferred historical decline of this species in secondary salinised systems. The use of non-lethal methods for ascertaining these tolerances would be ethically advisable, given the threatened status of the species (see for example Beatty et al. 2011). Climatic warming could also contribute to potential fish kills, as rising water temperatures can cause increased respiration rates and reduced dissolved oxygen concentrations that may fall below tolerance limits of aquatic biota (Bunn & Davies

1992; Davies 2010).

Positive initial steps have recently been taken towards conserving N. pygmaea with its listing under the Western Australian Wildlife Conservation Act 1950. This outcome should be viewed, not as an end in itself, but rather the beginning of an ongoing commitment by environmental managers responsible for the stewardship of the waterways that it occupies and the surrounding landscapes. This study has identified numerous knowledge gaps for this species including the need for further data on its reproduction, critical spawning habitats, longevity, growth and development, diet, disease, and particularly its salinity tolerance, as well as salinity tolerance of the early life-history stages of all fish species inhabiting salinised rivers in south-western Australia.

The study has also identified an important area for future research aimed at proactively managing rare fish species such as N. pygmaea and N. balstoni against climate change and other anthropogenic stressors. Crucial knowledge gaps on the characteristics of

160 artificial baseflow refuge habitats that house these rare species including bathymetry, habitat complexity, physico-chemistry, location, and hydrology need to be addressed, with a view towards the strategic creation of additional refuge habitats for these, and other, species. This is one potential strategy that can be used by scientists and environmental managers to confront the challenge of conserving the unique biodiversity that has evolved in this globally significant region.

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Chapter 6

General Conclusions

The depauperate but highly endemic freshwater fish fauna of south-western Australia has been, and continues to be, impacted by multiple anthropogenic stressors including secondary salinisation, river regulation, introduced species, and climate change (Morgan et al. 2003, 2004b; Beatty et al. 2011, 2013, 2014a, b; Morrongiello et al. 2011; Beatty &

Morgan 2013). Consequently, several species have suffered extensive range declines (e.g.

Morgan et al. 2014; Ogston et al. 2016) and the fauna now contains a number of species that are legislatively acknowledged as threatened (Table 1.1). In most cases, extensive land clearing, degradation of aquatic habitats, and resultant secondary salinisation, have been responsible for the loss of populations and fragmentation of distributions. Whilst, land clearing has eased since the late-1980s (Bradshaw 2012), other anthropogenic stressors have escalated in recent decades.

This thesis presents a body of research undertaken to investigate the likely influence of anthropogenic climate change on the impacts of instream barriers and the manner in which these structures are assessed and prioritised for works to mitigate their impacts on migratory fishes and aquatic ecosystems more broadly. South-western

Australia represents an ideal, if not unfortunate, model region for undertaking such an investigation due to the impact that climate change has already had in this region (as well as projected future climate change), the high prevalence of instream barriers, and its highly migratory and imperilled indigenous fish fauna. Included in this research, was an investigation into aspects of the life history of some of the region’s rarest and most endangered fish taxa, in order to ascertain the relative impact of instream barriers and other anthropogenic stressors and identify processes for managing these threats.

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6.1 Barriers to fish migration in a drying climatic region

As our knowledge of the life-cycles of the unique freshwater fishes of this region has expanded over the past 30 years, so too has our recognition of the importance of migration for these species (e.g. Beatty et al. 2011; 2014a). All of these species migrate for breeding purposes with the distances moved varying among species; ranging from tens of metres for L. salamandroides (Morgan et al. 2000) up to thousands of kilometres for diadromous species such as G. maculatus (Berra et al. 1995). Most of the region’s fish species, however, are recognised as potamodromous (Table 1.1), moving between permanent refuges and seasonally inundated habitats (e.g. lower order tributaries, swamps and wetlands) that are used as breeding grounds and nurseries (e.g. Pen & Potter 1991a, b; Beatty et al. 2014a). Migration within rivers is dependent on hydrological connectivity between source and destination habitats; however, barriers to migration such as dams, weirs and other river infrastructure can disrupt connectivity and impact migratory fishes.

Instream barriers also have the potential to delay the onset of flows which can provide cues for spawning in some fishes, thus potentially shortening the duration of the breeding season and reducing recruitment (Beatty et al. 2014b).

This study reviewed global trends in the construction and decommissioning of instream barriers and found that, in the developed world at least, dam removals have recently begun to outpace construction (WCD 2000; Pohl 2003; American Rivers 2010).

It is unclear if the same trend extends to small dams and other barrier types, due to a lack of reported data. In Australia, there have been few reported instances of instream barriers being decommissioned, but the construction of fishways to mitigate the impacts on migratory native fishes has increased in recent decades (e.g. Mallen-Cooper 1996).

Although only a small number of fishways have been installed in Western Australian rivers (e.g. Morgan & Beatty 2006; Beatty et al. 2007) and there are no reported barrier

163 removals, there is a growing recognition amongst environmental managers of the deleterious impacts of instream barriers on migratory fishes and the need to mitigate them

(Storer & Norton 2011; Norton & Storer 2011). This study, therefore, aimed to address this need by developing a process for identifying and prioritising instream barriers for removal or remediation (Figure 3.1).

This process incorporated a number of novel criteria designed to address the interactive effects of climate change on instream barrier impacts (both positive and negative), pertinent to non-perennial lotic systems in drying climatic regions such as south-western Australia; regarded as a model region for global climate change. Some of the unique aspects of this process are criteria that assess an instream barrier’s restriction of access to permanent refuge habitat, as well as the relative ecological value of permanent artificial refuge habitat provided by a barrier in its associated impoundment in light of projected climatic change (Table 3.1). A pilot study was used to trial the initial phases of the process (project budget was insufficient to undertake any barrier removal/remediation) in six different catchments and identified, assessed and prioritised over 60 potentially significant barriers to fish migration according to their impact on fish assemblages (Table 3.2).

This study revealed that low-altitude aerial surveys (using helicopter) are highly effective for validating information on artificial instream barriers gathered during the desktop review phase. More importantly, however, they allowed habitat features such as permanent refuge pools and natural barriers, which are crucial to understanding distributions and movements of fishes in intermittent systems but for which pre-existing data are mostly lacking in this region, to be identified and mapped.

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Only one of the top 10 ranked instream barriers in the prioritisation process appeared to be redundant, reflecting the typically high level of utility that such structures provide (e.g. stream gauging, vehicle crossings, water storage for irrigation). Therefore, an application of this process in its entirety in these catchments would likely result in remediation of barriers via fishway construction rather than decommissioning and complete physical removal.

6.2 Fishway use by critically endangered G. truttaceus during its breeding period

Remediation of instream barriers is most commonly achieved via the installation of fishways; of which there are relatively few in south-western Australia (but see Morgan &

Beatty 2006; Beatty et al. 2007, 2014b). One of these fishways was built specifically to facilitate access to habitat upstream of a gauging weir on the Goodga River for a critically endangered population of G. truttaceus (Morgan & Beatty 2006). Post-installation monitoring of the fishway proved that it was effective for this purpose (Morgan & Beatty

2006), but questions remained unresolved pertaining to patterns of use by individual fish, particularly during the autumn breeding season, which is atypically early cf. other fishes of the region (Morgan 2003).

Therefore, the current study aimed to address these questions by using an advanced biotelemetry method employing Passive Inducer Transponder (PIT) technology to track the movement of individual tagged G. truttaceus through the Goodga River

Fishway during the annual spawning period. This study was the first of its kind in the

State of Western Australia. The hypothesis that fishway movement is correlated with stream flow for this species in this system was confirmed. The most surprising result, however, was the low prevalence of upstream movement during the breeding season by

165 tagged adult fish (note that only adults were tagged). Morgan (2003) reported an upstream potamodromous breeding migration for this species in the Goodga River, therefore it was hypothesised that a high proportion of the tagged adult fish would be detected ascending the fishway during the spawning period. However, this proved not to be the case and was interpreted as evidence of site fidelity and home-ranging behaviour in this population. These results suggested that suitable spawning sites exist downstream of the weir, thus precluding the need for the majority of fish to ascend the fishway during a long-range potamodromous migration. This is encouraging from a management perspective, as the impact on spawning migrations of projected flow declines may be negligible for this population, although there could potentially be other impacts. For instance, lower flow might reduce access to inundated riparian spawning habitat, thus impacting recruitment. However, it appears that this species has a degree of biological plasticity as populations in the nearby Angove catchment were found to breed one month later than the Goodga population (P. Close, unpublished data) and differences in the seasonality of breeding have also been observed in Tasmanian populations of the species

(Humphries 1989). Thus, this species may be biologically well-equipped to contend with projected climate change.

Morgan and Beatty (2006) captured large numbers of juvenile G. truttaceus ascending the fishway during spring. Combined with the results of the current study, this suggests that the fishway provides greater ecological value in facilitating migration for recruitment and dispersal of juveniles, rather than spawning in this population of G. truttaceus. The observed expansion in range of G. truttaceus in the Goodga River is clearly attributable to the installation of the vertical-slot fishway (Morgan & Beatty 2006) and maintaining its functionality in the face of potential future flow declines associated with climate change will be crucial in the ongoing management of this critically

166 endangered population. Similar ecological benefits could also potentially be replicated in the neighbouring Angove River sub-population of G. truttaceus which is also impacted by significant fish barriers (dams and weirs).

6.3 Life-history of the Little Pygmy Perch

The Little Pygmy Perch (N. pygmaea) was only discovered in 2007, ironically during a baseline monitoring survey of the Mitchell River, which at the time, was under consideration for inter-basin abstraction of water to supplement declining levels in the nearby Quickup Dam in the Denmark River catchment (Morgan et al. 2008). This discovery, the first new teleost found in the region since 1978, resulted in these plans being shelved. Prior to the current study, almost nothing was known about this fish other than its exceedingly restricted range in the Hay River system and its close phylogenetic relationship to the sympatric congener N. vittata (Unmack et al. 2011; Morgan et al.

2013). One of the main aims of this study, therefore, was to elucidate its distribution, population size, and key life-history aspects including, age and growth characteristics, reproductive biology, and patterns of migration, as well as the main threats to its conservation.

Exhaustive targeted sampling in the Hay and Mitchell rivers revealed that the species is restricted to a handful of permanent refuge pools during baseflow and undertakes a short potamodromous migration into the lowermost reaches of the seasonally flowing Mitchell River. Significantly, sampling of nearby catchments resulted in the discovery of two previously unknown populations of this species in the Denmark and Kent catchments. An outlying fourth population was also discovered by chance in

Lake Smith, ~200 km west of the nearest population in the Kent River.

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The Hay/Mitchell population was sampled in most months over a consecutive two year period to elucidate life-history characteristics. With a few exceptions, sampling was intentionally non-destructive so as to minimise impact on this rare and threatened species and comply with Animal Ethics permits, and this had the unavoidable consequence of introducing some uncertainty in the determination of the biology. Nevertheless, some valuable results were obtained. Running ripe females were captured exclusively in the

Mitchell River from August to September. Examination of the gonads of a small sub- sample of spawning females revealed the presence of oocytes at various stages of development, indicating that this species is a serial spawner, much like its congener N. vittata (Pen & Potter 1991a). Examination of the otoliths of these fish revealed that maturity is likely reached as fish approach the end of their second year, i.e. one year later than N. vittata (Pen & Potter 1991a). This, and the fact that N. vittata are capable of spawning in the Mitchell River as well as the salinised Hay River main channel, appears to explain why N. vittata dominate numerically over N. pygmaea and N. balstoni in this system.

The Denmark River population predominantly occurred in one tributary in the upper catchment in a series of natural and artificial refuges (i.e. fire-management ponds).

Fish found in this system were found to grow to a much larger size than those captured in the Hay and Kent rivers. The age of these large specimens was unknown as otoliths were not examined; however, there is a possibility that fish in this non-salinised tributary are growing faster and/or living longer than conspecifics in the salinised Hay and Kent catchments. Salinity has been shown to impact growth rates of other freshwater fishes

(e.g. Likongwe et al. 1996; Küçük 2013).

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The quantity and quality of natural refuge habitat will likely decline in the future due to climate change, endangering the faunal communities that depend on these habitats.

The discovery of N. pygmaea in artificial refuge habitats provides an example, albeit unintentionally, of a potential means of offsetting losses of natural refuges. Research is urgently required into the efficacy of artificial refuge provision (including barrier- associated impoundments) as a strategy for conserving aquatic biodiversity (particularly for threatened taxa such as N. pygmaea) in south-western Australia. Regardless of the expansion of the known range, this species qualified as Endangered under IUCN Red List assessment criteria (IUCN 2014) and has recently been listed as such on Schedule 1 of the Western Australian Government’s Wildlife Conservation Act 1950. It is one of the most endangered fishes in the region and urgently requires further study and proactive management.

6.4 Interactions between climate change and key anthropogenic stressors of freshwater fishes in south-western Australia

Besides the direct impacts that projected climate change will have on aquatic ecosystems and migratory fishes of south-western Australia, such as reducing the quantity and quality of available permanent refuges and seasonally inundated habitats, exacerbation of the impacts of other anthropogenic stressors is also likely. An obvious example is the increase in demand for potable water in the future as the region’s population grows

(CSIRO 2009a, b). Meeting this demand will place further strain on already stressed aquatic ecosystems through increased abstraction of surface and groundwater resources.

There has been a ~50% reduction in the amount of inflow to the region’s major water storage reservoirs since the mid-1970s, which now contribute < 40% of water consumed

169 in the region’s major cities (Petrone et al. 2010). By 2060, it is expected that these dams will no longer provide a reliable contribution to the water supply due to climatic drying

(Water Corporation 2009; SoE 2011). Subsurface water resources (e.g. Yarragadee

Aquifer) will therefore be increasingly relied upon to meet future demands, and the combined effects of abstraction and projected decline in groundwater recharge due to climate change (e.g. Del Borrello 2008; CSIRO 2009b) will place unprecedented hydrological stress on aquatic communities, particularly in systems impacted by secondary salinisation (e.g. Beatty et al. 2011, 2014a). Whilst the utility of large storage dams may decrease in the future due to climate change, it is expected that the number of small private dams will continue to increase in order to meet future anthropogenic water demands in rural areas (see Chapters 2 & 3). The impacts of instream barriers on aquatic ecosystems and migratory fishes (e.g. disruption of habitat connectivity; thermal stratification in impoundments) will be exacerbated by flow reduction due to anthropogenic climate change (see Chapter 2).

Many of the alien fish species that have established self-maintaining populations in the region are ‘climatically mismatched’, i.e. they originate from non-Mediterranean climatic regions (Beatty & Morgan 2013). Higher future temperatures due to climate warming may therefore serve to facilitate further alien species introductions and provide increasing competitive advantages to species native to tropical and sub-tropical regions over the indigenous regional fauna (Morrongiello et al. 2011; Beatty & Morgan 2013).

One such species of particular concern is Gambusia holbrooki, which is currently the most widespread alien fish species in Australia (Morgan et al. 2004). Its well-known deleterious effects (Gill et al. 1999) have the potential to increasingly impact upon populations of small-bodied indigenous species (Morrongiello et al. 2011).

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The effect that climate change will have on secondary salinisation in the region in the future is uncertain (Mayer et al. 2005); however, it is expected that salinity levels in salinised systems will become elevated as the dilution effect of streamflow is reduced due to future rainfall declines (Cañedo-Argüelles et al. 2013). Unseasonal summer rainfall events, which are increasing in frequency and intensity in salt-affected parts of the region

(BoM 2015), may also contribute to elevated salinity through the flush of highly concentrated surface salt deposits in runoff (see Chapter 5). Declining groundwater

(CSIRO 2009b; Barron et al. 2012) will reduce freshwater inputs in secondary salinised systems such as the Blackwood, thus threatening the viability of refuge habitats used by imperilled stenohaline species in this catchment such as Nannatherina balstoni and

Galaxiella munda (Beatty et al. 2011; 2014a).

6.5 Recommended areas for further investigation

 The establishment of a coordinated and rigorously verified global inventory of

instream barriers, whilst a huge undertaking, would aid tremendously in strategic

planning decisions around the mitigation of instream barrier impacts. The use of

prioritisation processes that account for the influence of climate change on barrier

impacts, such as that developed in this study, should also be advocated in other

global regions.

 Large water storage reservoirs currently contribute a significant amount to the

regional water supply and some are also used to supplement reduced flow rates

through environmental water releases (e-flows), but the contribution is projected

to diminish to negligible levels in the future due to rainfall and surface flow

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declines. Assuming that such barriers are not identified by the prioritisation

process developed during this study as a high priority for removal, their

importance as a source of e-flows to sustain important downstream refuge habitats

is likely to increase. Furthermore, there is potential for reservoirs to become

important artificial refuges; however, many currently offer little in the way of

suitable habitat for native fauna. Habitat rehabilitation (e.g. re-snagging of banks

to increase habitat complexity; revegetation of riparian zones to provide important

shade; feral fish control) could be trialled to ascertain their suitability for

sustaining translocated populations of threatened fishes. Certainly, the occurrence

of N. pygmaea and N. balstoni in natural lentic habitats such as Lake Smith

suggests that at least some threatened fish species might be able to adapt to large

artificial lentic habitats.

 The use of aerial survey via helicopter proved highly effective for determining the

spatial distribution of migration barriers and refuge habitats in non-perennial

rivers. Further survey work should be undertaken to continue filling this crucial

knowledge gap in other non-perennial systems housing threatened species.

Ongoing advancements in remote sensing technologies that can detect refuge

pools at fine spatial scales (e.g. Lidar) show much promise and these methods

should be trialled in the region and compared with manual aerial survey methods.

 Filling knowledge gaps in the spatial distribution of the regional fish fauna is

crucial as this information is vital when making informed management decisions,

such as in the application of the barrier prioritisation process developed in this

study. Three new sub-populations of N. pygmaea were discovered during this

study, clearly demonstrating the value in targeting previously unsurveyed sites or

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those that have been poorly surveyed for fishes. Environmental DNA techniques

are showing much promise for the detection of freshwater fishes and could be

used to screen waterways for previously undetected populations. Regular

monitoring of sites housing threatened species is also vital as it enables declines in

population size and/or habitat condition to be detected in a timely manner so that

emerging threats can be identified (e.g. riparian zone degradation, water quality

issues, novel species introductions, salinisation) and addressed, if possible.

 Accurate hydrological modelling of surface and sub-surface water resources is

crucial in order to monitor the impacts of climate change and water abstraction on

water availability and flow around critical habitats such as seasonal spawning

sites, refuge pools, and at migration barriers. Accurate projections of hydrological

impacts of climate change on such habitats is vital knowledge that can feed into

the barrier prioritisation process and guide management decisions surrounding

barriers and their impoundments (see Chapter 2).

 Completion of the barrier identification and prioritisation process that was

commenced in the pilot study could be valuable and provide a template for future

barrier mitigation projects. This would allow any potential shortcomings of those

phases that were not trialled in the current study (e.g. stakeholder engagement,

barrier removal/remediation, pre- and post-mitigation monitoring) to be identified

and modified in order to improve the process, and accurate financial costings of a

full application of the process could also be determined. An application of the

process in its entirety in other rivers of south-western Australia that house

threatened migratory fishes would also be valuable.

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 The study of movement patterns of G. truttaceus through the Goodga River

Fishway should be continued in order to obtain a more comprehensive temporal

dataset that will allow any seasonal or inter-annual variation in movement patterns

to be ascertained.

 The biological study of N. pygmaea was constrained by the ethical requirements

to minimise destructive sampling of this rare and endangered fish. As a result, a

great deal still remains to be learned about this species including: physiological

tolerances to temperature and salinity (including early life history stages of this

and all south-western Australian freshwater fishes); dietary analysis; microhabitat

requirements; ontogeny/larval development; reproductive biology and age and

growth characteristics (which are currently only partially known); predators and

trophic position; interactions with sympatric alien species (G. holbrooki);

diseases/parasites; phylogenetics of the Lake Smith sub-population; and further

targeted sampling in the intervening area between Lake Smith and the Kent River

to search for previously undetected populations.

 Artificial refuges (fire management ponds) are currently utilised by N. pygmaea,

N. balstoni and other native teleosts in the Denmark River catchment, thus there

appears to be enormous potential for mitigating future population declines or

losses of this species through the provision of additional artificial refuges or intra-

basin translocation into existing artificial refuges that are currently unoccupied by

these species. Further research into the efficacy of this proactive conservation

strategy should be given top priority as it has enormous potential to offset future

losses of natural refuge pools caused by climate change and water abstractions.

174

 A captive breeding program should also be instigated for N. pygmaea (possibly at

the population level) to help mitigate potential catastrophic losses, particularly for

the highly vulnerable Hay River sub-population. The use of captive bred stock in

future studies could also help to circumvent potential limitations on the numbers

of individuals of this endangered species that can be destructively sampled in the

pursuit of further knowledge of important aspects of the biology of the species in

order to better manage its conservation.

6.6 Closing remarks

The projected scale and pace of future climatic changes in the region is a daunting prospect not only for biodiversity management, but also for human societies and economies. One of the major environmental challenges faced globally is finding a balance between human water demands and those of the environment. Many of the aquatic ecosystems in south-western Australia are degraded, while the areas that remain near pristine face a number of serious threats, including artificial instream barriers. This study has revealed how climate change will, on the one hand, exacerbate negative ecological (e.g. hydrological disconnectivity) and socio-economical (e.g. reduced cost- efficiency of maintaining storage of diminishing surface water resources) impacts of instream barriers, while on the other hand, amplifying the potential positive impacts (e.g. providing artificial refuge for threatened species). Given the recent declines of many freshwater fishes in the region, it is crucial that instream barriers and other key threatening processes are addressed and that monitoring of these species and their habitats continues. To this end, a rigorous process for identifying, assessing and prioritising artificial instream barriers for consideration of removal or remediation was

175 developed in this study that could be applied in drying climatic regions characterised by non-perennial lotic systems such as south-western Australia. This thesis adds to the body of research that has contributed to our understanding of how best to monitor, rehabilitate, and manage the aquatic ecosystems of this region, and it is vital that further work continues in this field to expand our collective knowledge. Such work must also become more efficient and focussed in the future in the face of rapidly escalating threats and diminishing expenditure on environmental management in order to deliver the maximum possible benefit to the unique aquatic biodiversity of south-western Australia.

176

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Appendix 1. Information flyer and questionnaire mailed to riparian landholders.

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Appendix 1. (cont.) Information flyer and questionnaire mailed to riparian landholders.

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Appendix 1. (cont.) Information flyer and questionnaire mailed to riparian landholders.

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Appendix 2. Instream barrier aerial survey data sheet