Utility of a multi-faceted approach in determining the habitat use of endangered euryhaline elasmobranchs in a remote region of northern Australia
Jeff M. Whitty
This thesis is presented for the degree of Master of Philosophy of Murdoch University
2011
DECLARATION
I declare that this thesis is my own account of my research and contains as its main
content work that has not previously been submitted for a degree at any tertiary
education institution.
…………………………………………………
Jeff M. Whitty
2011
Frontispiece: P. microdon and G. garricki photos by Dean Thorburn
2 Abstract
The overriding aim of this thesis was to explore the habitat use of the critically endangered freshwater sawfish (Pristis microdon) and northern river shark (Glyphis garricki) in the remote Kimberley region of northern Western Australia. This information has been largely lacking and is critical for the management and conservation of these species. Habitat use of these species was documented through the use of long-term catch, environmental and tagging (conventional, acoustic and satellite) data, which was acquired between 2005 and 2009 in the Fitzroy River and
King Sound, Western Australia. Monitoring of the various environmental parameters demonstrated that the study area is extremely dynamic, with significant seasonal changes in abiotic variables such as water flow, depth, temperature, turbidity and salinity. Catch data demonstrated that juvenile P. microdon inhabit the Fitzroy River for three to five years, at which time individuals begin to emigrate from the river, prior to maturation. Catch data also demonstrated that juvenile and adult male G. garricki inhabit highly turbid nearshore waters throughout the upper King Sound.
Relative abundance of P. microdon in the river varied seasonally and annually, and differed between size classes. A decrease in relative abundance between the early and late dry season, which was only significant with young of the year P. microdon, was attributed to mortality as well as dispersal of individuals. Catches of G. garricki were rare, although this species was relatively abundant in highly turbid waters in King
Sound. Foraging, depth use and inter-pool movements of P. microdon in the Fitzroy
River were influenced by depth, flow and light intensity/turbidity and potentially by salinity and/or water temperature. However, the response to environmental variables differed between P. microdon size classes, possibly due to differences in trophic and
3 habitat requirements of the various size classes. Results from this study demonstrated that P. microdon is sensitive to environmental change, but appear to endure/recover from short-term (months) negative impacts through behavioural regulation. It is not possible at this time to positively conclude about the impacts of specific environmental changes on G. garricki habitat use, due to the limited data from G. garricki captures and tagging. However, as all G. garricki captured in this study were observed to inhabit tidal creeks and river mouth areas, the destruction of such areas or restriction to and from such areas is likely to negatively impact this species.
4 Acknowledgments
I am deeply grateful to a number of people and organizations who were of great value in this research. From these people I have learned the meaning of hard work and generosity. First and foremost, I would like to thank my supervisors Dr David
Morgan and Dr Colin Simpfendorfer. I would like to thank Dr Morgan for giving this yank the opportunity to become a part of Team Sawfish, for providing me with the knowledge and support to make this research possible and for acting as a great PR manager. I would like to thank Dr Simpfendorfer for his guidance, knowledge and for his always insightful reviews, which greatly improved this thesis. I am also greatly indebted to Dr Dean Thorburn, Stirling Peverell, Dr Brendan Ebner and Dr Stephen
Beatty for going well out of their way to help with this project in and out of the field, and for helping me to see the big picture. I would also like to express my extreme gratitude to the countless people who were brave enough to face crocodiles, mozzie attacks, sweltering heat and no sleep to help with the field work, including Mark
Allen, Ryan Bell, Brad Norman, Nicole Phillips, Simon Visser as well as the Yiriman
Project and the Jarlmadangah and Kimberley Land Council Rangers, which consisted of Nyaburu Watson, Kimberley Watson, Cannie Watson, Jon Watson, Josh Albert,
Angus Butt, William Lennard, Neville Poelina, Simon Keenan, Will Bennett, Darryl
Combs, Mick Apanah and Hugh-Wallace Smith as well as Travis Fazeldean to whom
I would like to especially thank for his hard work and patience while working in the
King Sound. I would also like to convey my appreciation to Arthur and Leena as well as the countless volunteers of Broome and Derby for their help and support. Special thanks go to Jim and Gerry Kelly for their support, friendship as well as for the two
5 walls and a roof provided during field work. A special thank you also goes out to
Ferdy Bergmann and to his crew and family for their overwhelming generosity and knowledge given in the interest of conserving the fishes of the King Sound.
I am also grateful to everyone at Murdoch University for their support and feedback. Thank you to Dr James “Cobra” Tweedley, Nicole Phillips and Michelle
Gardner for helping me maintain my sanity. Thank you to Dr Alan Lymbery and again to Dr Tweedley and Dr Beatty for their patience and assistance with statistical analyses.
I would also like to thank the Department of Environment, Water, Heritage and the Arts and the Western Australian Government’s State NRM for funding this project. Additionally, I owe my gratitude to the New South Wales Aquarium Society,
Dr Zeb Hogan and National Geographic, Mitre 10 Derby and Dr Simon de Lestang for their generous contributions to this study. Also, I am grateful to the Australian
Broadcast Corporation, National Geographic, Animal Planet and associated crews for allowing us to share this work with the world.
I would like to express my deep gratitude to Dr Heidi Dewar for all the great advice and support through the years. It was her advice and support to go see the world that started this great adventure. I would also like to thank Sue Donohue Smith for her guidance and friendship through the years. It was Sue that helped spark my interest in marine biology and helped me along my career path.
Lastly, I would like to thank my friends and family for their support and interest through the years. I would like to express my appreciation to the SD crew who always welcomed me back with open arms and a warm couch. I thank my
Grandma and Mary-Ann for their countless emails and numerous prayers. A special thank you goes to Shannon, Mark, Jake and Zach for their support and love. Finally, I
6 would like to express my deepest gratitude to my mom and dad. It has been a long adventure, which their love, support and patience has made possible. Most importantly, it was their encouragement to follow my dreams that gave me the confidence and ability to complete this important milestone.
7 Table of contents
Abstract ...... 3 Acknowledgements ...... 5 Table of contents ...... 8 List of tables ...... 11 List of figures ...... 12
Chapter 1: General Introduction ...... 15 1.1 Habitat use of elasmobranchs ...... 15 1.2 Tagging and tracking methodologies ...... 18 1.3 Pristis microdon and Glyphis garricki ...... 21 1.4 Status and threats to Pristis microdon and Glyphis garricki ...... 24 1.5 Aims of the study ...... 26
Chapter 2: Environments of the Fitzroy River and King Sound, Western Australia ...... 29 2.1 Introduction ...... 29 2.2 Materials and methods ...... 34 2.2.1 Study sites ...... 34 2.2.2 Depth ...... 39 2.2.3 Temperature ...... 39 2.2.4 Salinity ...... 40 2.2.5 Turbidity and light attenuation ...... 41 2.2.6 Dissolved oxygen and pH ...... 41 2.2.7 Tidal cycle ...... 42 2.2.8 Data analysis ...... 43 2.3 Results ...... 44 2.3.1 Depth ...... 44 2.3.2 Temperature ...... 47 2.3.3 Salinity ...... 54 2.3.4 Turbidity and light attenuation ...... 57 2.3.5 Dissolved oxygen and pH ...... 59 2.3.6 Tidal cycle ...... 60 2.4 Discussion ...... 60 2.4.1 Temperature ...... 61 2.4.2 Salinity ...... 63 2.4.3 Turbidity and light attenuation ...... 64 2.4.4 Dissolved oxygen and pH ...... 66 2.4.5 Conclusion ...... 67
Chapter 3: Population demography and life history of Pristis microdon and Glyphis garricki in the Fitzroy River and King Sound, Western Australia ...... 70 3.1 Introduction ...... 70 3.2 Material and methods ...... 72 3.2.1 Study site ...... 72
8 3.2.2 Sampling ...... 74 3.2.3 Data analysis ...... 76 3.3. Results ...... 80 3.3.1 Pristis microdon ...... 80 3.3.1.1 Catch composition and distribution ...... 80 3.3.1.2 Catch per unit effort ...... 85 3.3.1.3 Growth ...... 87 3.3.1.4 Mortality and survivorship ...... 90 3.3.2 Glyphis garricki ...... 91 3.4 Discussion ...... 94 3.4.1 Distribution ...... 95 3.4.2 Population composition ...... 97 3.4.3 Population life history ...... 100 3.4.3.1 Growth ...... 100 3.4.3.2 Mortality ...... 101 3.4.4 Catch per unit effort ...... 102 3.4.5 Conclusion ...... 105
Chapter 4: Depth use and inter-pool movement of juvenile Pristis microdon in the Fitzroy River ...... 107 4.1 Introduction ...... 107 4.2 Materials and methods ...... 108 4.2.1 Study site ...... 108 4.2.2 Acoustic array and range testing ...... 110 4.2.3 Tagging ...... 113 4.2.4 Environmental conditions and prey abundance ...... 115 4.2.5 Data analysis ...... 116 4.3 Results ...... 117 4.3.1 Acoustic array and range testing ...... 117 4.3.2 Acoustic tag performance ...... 118 4.3.3 Depth use ...... 121 4.3.4 Inter-pool movement ...... 126 4.3.5 Relative abundance of Pristis microdon prey ...... 127 4.4 Discussion ...... 128 4.4.1 Depth use ...... 129 4.4.2 Inter-pool movement ...... 131 4.4.3 Acoustic tag performance and utility ...... 134 4.4.4 Conclusion ...... 136
Chapter 5: Utility of conventional and satellite tags in the monitoring of large-scale movement and habitat use patterns of Pristis microdon and Glyphis garricki ...... 138 5.1 Introduction ...... 138 5.2 Materials and methods ...... 140 5.2.1 Study site ...... 140 5.2.2 Tagging ...... 142 5.2.3 Data analysis ...... 146 5.3 Results ...... 147 5.3.1 Rototags ...... 147
9 5.3.2 Satellite tags ...... 150 5.4 Discussion ...... 152 5.4.1 Rototags ...... 153 5.4.1.1 Utility ...... 153 5.4.1.2 Habitat use ...... 155 5.4.2 Satellite tags ...... 157 5.4.2.1 Utility ...... 157 5.4.2.2 Habitat use ...... 158 5.4.3 Conclusion ...... 159
Chapter 6: General discussion ...... 161 6.1 Pristis microdon habitat use ...... 161 6.2 Glyphis garricki habitat use ...... 163 6.3 Potential impacts of habitat alterations on the habitat use of Pristis microdon and Glyphis garricki...... 164 6.4 Future research ...... 167
References ...... 170
Appendix I: Pristis microdon catch data (2005 to 2009) ...... 194
10 List of tables
Table 2.1. Minimum and maximum temperatures of monitored pools in 2007 to 2009 ...... 48
Table 2.2. Correlation coefficients (p < 0.05, score not given if p > 0.05) of river water and air temperatures, as well as river discharge of monitored pools at varying temporal and spatial scales ...... 52
Table 2.3. Seasonally separated correlation coefficients (p < 0.05, score not given if p > 0.05) of river water and air temperature, river discharge and tidal height at varying spatial scales employing daily and weekly means ...... 53
Table 3.1. Percentage of yearly 20-m net sets for each respective gillnet mesh size deployed ...... 75
Table 3.2. Methods used for calculating natural mortality (M) estimates . . . . . 80
Table 3.3. Number of captured male/female Pristis microdon for respective years and size classes in the Fitzroy River ...... 84
Table. 3.4. Total CPUE (Pristis microdon 100 h-1 of 20-m net set) for YOY and >0+ juvenile Pristis microdon for the 2002 to 2009 early and late dry seasons, in the Fitzroy River ...... 86
Table 3.5. Growth data of Pristis microdon at liberty for >1 month in the Fitzroy River ...... 88
Table 3.6. Scores and von Bertalanffy growth function parameters produced from ELEFAN and PROJMAT methods using non-seasonal and seasonal growth curves ...... 89
Table 3.7. Natural mortality (M) and annual survivorship (S) of juvenile Pristis microdon ...... 91
Table 3.8. Glyphis garricki captures in 2007 and 2008 in King Sound ...... 93
Table 3.9. Number of elasmobranchs, by species, captured within the upper King Sound (excludes the NW King Sound area) ...... 94
Table 4.1. Acoustic tag deployment information for 2007 to 2009 ...... 114
Table 4.2. Detection details of acoustic tags deployed in 2007 to 2009 ...... 119
Table 5.1. Satellite tag deployment information ...... 145
Table 5.2. Pristis microdon recaptures in the Fitzroy River in 2005 to 2010 . . 149
11 List of figures
Fig. 2.1. (a) Map of sampled sites in the Fitzroy River (2007 to 2009), and satellite photos of (b) estuarine and (c) freshwater pools in the dry season at similar magnification ...... 36
Fig. 2.2. Sampled sites (red dots) and areas (shaded) within the King Sound (2007 to 2009). Green shading = NW King Sound area, orange shading = SW King Sound area, yellow shading = SE King Sound area, red shading = Point Torment area, blue shading = Stokes Bay area ...... 37
Fig. 2.3. Maximum daily stage height (m) at Camballin Barrage station, Fitzroy River for the 2002 to 2009 wet seasons; data courtesy of the Department of Water, Government of Western Australia . . . 45
Fig. 2.4. Maximum daily discharge (m3 s-1) at Willare station, Fitzroy River for the 2002 to 2009 wet seasons; data courtesy of the Department of Water Government of Western Australia . . . 45
Fig. 2.5. Total annual (December to November) river discharge (ML) at Willare station, Fitzroy River for 2002 to 2009; data courtesy of the Department of Water, Government of Western Australia . . . 46
Fig. 2.6. Depth profiles of (a) estuarine and (b) freshwater pools. Depth profiles extend to the maximum detection range up (+) and downstream (-) of moored acoustic receivers (see Chapter 4) . . . . . 47
Fig. 2.7. Recorded air, King Sound and Fitzroy River (Milli Milli = estuarine representative, Myroodah Crossing = freshwater representative) temperatures between June 2007 and October 2009; air temperature data was courtesy of the Bureau of Meteorology, Commonwealth of Australia (2010) ...... 49
Fig. 2.8. Monthly mean (± s.e) diel temperature ranges of monitored Fitzroy River pools ...... 50
Fig. 2.9. Tide and water temperature (ºC) profiles in Milli Milli in the late dry season 2008 ...... 51
Fig. 2.10. Mean (+ s.e.) salinity of monitored sites in nearshore waters of the Fitzroy River and King Sound in (a) June/July and (b) September to November ...... 56
Fig. 2.11. Light intensity (lux) in Telegraph Pool (0.5 m) and Myroodah Crossing (1.5 m) at the start of the 2009 wet season ...... 58
12 Fig. 2.12. Logged light intensity (lux) in Myroodah Crossing (at a depth of 1.5 m) during the 2009 wet season, in relation to the stage height recorded at the Camballin Barrage (~38 km upstream of Myroodah Crossing) ...... 58
Fig. 3.1. Sampled sites (red dots) and study areas (shaded) in King Sound and Fitzroy River between 2005 and 2009. Green shading = NW King Sound area, orange shading = SW King Sound area, yellow shading = SE King Sound area, red shading = Point Torment area, blue shading = Stokes Bay area ...... 73
Fig. 3.2. Pristis microdon capture locations in 2005 to 2009 (red dot), and in 2002 to 2004 (Thorburn et al. 2007) (blue square) ...... 82
Fig. 3.3. Pristis microdon total length-frequency histograms for 2002 to 2009 ...... 83
Fig. 3.4. Total Pristis microdon CPUE (circle) and percentage of positive Pristis microdon gillnet captures (square) in the early (June to July; black symbol) and late (October to November; gray symbol) dry season (2002 to 2009); only freshwater pools were sampled in 2005 as well as late 2004 and late 2009 ...... 87
Fig. 3.5. Best fit growth curves produced by non-seasonal (a) ELEFAN and (b) PROJMAT methods for juvenile Pristis microdon plotted on original length-frequency data ...... 88
Fig. 3.6. Age-dependent indirect natural mortality estimates of juvenile Pristis microdon calculated using Peterson and Wrobleski (1984), Chen and Watanabe (1989) and Lorenzen (1996) methods ...... 91
Fig. 3.7. Glyphis garricki capture locations in King Sound in 2007 and 2008 (red dot), and in 2002 to 2004 (blue square) (Thorburn and Morgan 2004) ...... 92
Fig. 4.1. Acoustic monitoring stations (red dots) in the Fitzroy River, Western Australia. Crs. = Crossing, Lwr = Lower, Pl. = Pool, Upr = Upper ...... 110
Fig. 4.2. Activity of (top) acoustic tags and (bottom) hydrophones in the respective pools. Tag deployments are represented by triangles and detected transmissions are represented by circles. 1: Milli Milli, 2: Snag Pool, 3: Telegraph Pool, 4: Langi Crossing, 5: Myroodah Crossing 1, 6: Myroodah Crossing 2, 7: Camballin Pool, 8: Lower Barrage Pool, 9: Upper Barrage Pool ...... 120
13 Fig. 4.3. Depth profiles of Pristis microdon (a) <1000 mm, (b) 1400 to 1650 mm and (c) >1850 mm TL in the dry season. Only tags with >30 hits per hour for every hour are displayed. Note that the scales of the x and y-axes vary between graphs ...... 123
Fig. 4.4. Recorded depths of Pristis microdon tag (a) 17 (Camballin Pool), (b) 18 (estuarine pools), (c) 19 (estuarine pools), (d) 9 (estuarine pools) and (e) 16 (Myroodah Crossing), in comparison to river stage height in the 2009 wet season. No other tags were detected during the wet season ...... 124
Fig. 4.5. Mean hourly depths (± s.e.) of (a) <1000 mm TL, (b) 1400 to 1650 mm TL and (c) >1800 mm TL Pristis microdon in the Fitzroy River ...... 125
Fig. 5.1. Survey sites in the Fitzroy River and King Sound, Western Australia ...... 142
Fig. 5.2. (a) Rototag and (b) SPOT-5 satellite tag attachment sites and methods ...... 146
Fig. 5.3. Initial capture (diamond) and recapture (circle) sites of conventionally tagged Pristis microdon observed to move between pools in the Fitzroy River...... 150
Fig. 5.4. Deployment (diamond) and calculated transmission locations (circles) for tagged Pristis microdon and Glyphis garricki in King Sound ...... 152
14 Chapter 1: General introduction
Successful protection of a species is dependent upon the understanding of its habitat requirements and use (Murphy and Noon 1992; Lambeck 1997). Hall et al. (1997) define habitat use as “the way an animal uses (or ‘consumes’, in a generic sense) a collection of physical and biological components”. As such, a change in these components is likely to lead to changes in habitat use. By observing temporal and spatial movement patterns of an individual, acquirable through tagging and/or catch data, in relation to changes in abiotic and biotic variables, we can better understand the habitat requirements and habitat use of an animal within its home range
(Morrissey and Gruber 1993). Knowledge of these interactions is invaluable in understanding how outside influences can impact a species. It is also essential in designing and implementing effective management strategies. This is directly applicable to critically endangered species, such as the freshwater sawfish (Pristis microdon Latham, 1794) and northern river shark (Glyphis garricki Compagno, White and Last 2008), formerly known as Glyphis sp. C, (sensu Compagno and Niem 1998), which face ongoing threats of fishing and habitat degradation (Pogonoski et al. 2002;
Stevens et al. 2005).
1.1 Habitat use of elasmobranchs
Behavioural ecology of elasmobranchs has been the focus of numerous studies, with a growing attention to habitat use and movements (e.g. Morrissey and Gruber 1993;
Sundström et al. 2001; Cartamil et al. 2003; Heupel et al. 2004; Peverell 2005;
Heithaus et al. 2006; Sims et al. 2006; Dewar et al. 2008; Carlisle and Starr 2009).
Such studies have shown that an organism will use various components of a habitat in
15 a way that will maximise its potential fitness (MacArthur and Pianka 1966; Werner et al. 1983; Hugie and Dill 1994; Krause et al. 1998; Sims 2003). Often a single microhabitat (defined by this study as: an area within an animal’s daily habitat range that is distinct from surrounding areas in regards to the resources used by the species in question, e.g. bottom and surface waters of riverine pools), which may be used for foraging, cover, reproduction and/or other life history traits, will not contain all of the resources needed by a species. Consequently, individuals may be found to move between microhabitats frequently. An example of this is the movement of various elasmobranchs between foraging grounds and other microhabitats for thermoregulation (Matern et al. 2000) or sanctuary (Holland et al. 1993; Cartamill et al. 2003). Habitat use patterns and selection may also change as the availability and accessibility of resources as well as the needs of an animal change with growth
(Bethea et al. 2004), season (Smith and Merriner 1987; Lucifora et al. 2005; Carlisle and Starr 2009) and/or during specific life moments, such as for copulation or pupping (Pratt and Carrier 2001). When a habitat can no longer benefit an organism the individual will likely select a more suitable habitat, as seen in the movement of sharks from nursery areas to deeper waters with growth (Pikitch et al. 2005; Duncan and Holland 2006).
Within each habitat the behaviour of an animal is influenced by numerous abiotic and biotic factors. Abiotic factors, such as temperature (Matern et al. 2000;
Sims et al. 2006), salinity (Heupel and Simpfendorfer 2008), tides (Medved and
Marshall 1983; Smith and Merriner 1985; Ackerman et al. 2000; Carlisle and Starr
2009) and turbidity/light intensity (Miner and Stein 1996; Cartamil et al. 2003;
Scheuerell and Schindler 2003; Dewar et al. 2008; Chiu and Abrahams 2010) have been shown to influence habitat selection and use of numerous elasmobranch species.
16 Tides for example, have been shown to influence the swimming direction and foraging efforts of many nearshore elasmobranchs (Medved and Marshall 1983;
Smith and Merriner 1985; Ackerman et al. 2000; Carlisle and Starr 2009). Interaction between organisms (i.e. predation and competition) is a biological factor that has been shown to influence habitat use (MacArthur and Levins 1964; Schlosser 1987; Hugie and Dill 1994; Lucifora et al. 2005; Heithaus and Dill 2006; Papastamatiou et al.
2006; Torres et al. 2006; Fischhoff et al. 2007). The effect of this has been shown to give rise to habitat partitioning (Schoener 1968; Schoener 1974; Papastamatiou et al.
2006) and cause less dominant or fit organisms to temporarily take refuge in less productive areas. This behaviour can result in decreasing net energy gain but also reduces competitive pressures and predation risk (MacArthur and Levins 1964;
Schlosser 1987; Lucifora et al. 2005; Heithaus and Dill 2006; Fischhoff et al. 2007).
In such cases, species are generally seen to move back to productive areas when resources cannot support them (Heithaus et al. 2007a) or risk of predation decreases
(by absence of predators or increased growth/fitness) (Lucifora et al. 2005; Wirsing et al. 2007; Kerford et al. 2008). Having such a profound effect on habitat use and selection, it is important to consider both abiotic and biotic variables in habitat use analyses.
Although the understanding of elasmobranch movements and habitat use is increasing, there continues to be an absence of knowledge for many species found in remote areas, such as northern Australia. A relatively undeveloped, harsh and extreme area, very little research has been conducted on elasmobranchs in this area, with less attention given to the behavioural ecology of these fishes. It is hypothesised that the extreme fluctuations of environmental variables in areas like the Fitzroy River
17 and King Sound, Western Australia, are likely to influence the fitness and behaviours of elasmobranchs in these areas.
1.2 Tagging and tracking methodologies
Tagging of fishes has been a valuable method used for hundreds of years to explore fish movement and habitat use in marine and freshwater environments (Kohler and
Turner 2001; Voegeli et al. 2001). Tagging/marking of fish began with simple body markings (e.g. fin clips), and progressed through the years, giving way to conventional and eventually electronic tags (Kohler and Turner 2001; Arnold and
Dewar 2001; Voegeli et al. 2001). Both conventional and electronic methods have been shown to be highly useful with elasmobranchs (Klimley et al. 2001; Voegeli et al. 2001; Heupel et al. 2004; Collins et al. 2007; Weng et al. 2008; Jorgensen et al.
2009; Skomal et al. 2009). However, the effectiveness of a tag type in a study depends on factors such as the species under investigation, study area, financial budget and aims of the study, as each tag type is accompanied with its own advantages and disadvantages.
Individually marked conventional tags, used in mark and recapture studies, allow researchers to monitor large scale movements, growth and population statistics
(Kohler and Turner 2001). Mark and recapture studies are dependent upon the recapture of tagged individuals but often have a low rate of return (Kohler and Turner
2001). Studies often mark and release a large number of animals in order to recapture enough fish for meaningful statistical analyses (Thorson 1971; Stevens et al. 2000;
Kohler and Turner 2001; Kohler et al. 2002). Various conventional tag types have been used in mark and recapture studies, including the more common Rototags,
Petersen disc tags and spaghetti tags (Kohler and Turner 2001). Rototags in general
18 have been found to have longer retention times than Petersen disc tags (Thorson 1971;
Xiao 1999; Kohler and Turner 2001), and a lower risk of premature shedding (Smith and McPherson 1981; Kohler and Turner 2001) and incur less damage to surrounding tissue than spaghetti tags (Heupel et al. 1998). As such, they are arguably one of the more reliable conventional tags used for elasmobranch mark and recapture studies.
However, this is dependent upon the species and sizes of individuals involved.
More advanced electronic tags (e.g. acoustic transmitters, radio transmitters, archival tags and satellite tags) most often do not require recapture of animals (Lucas and Baras 2000; Arnold and Dewar 2001; Ebner et al. 2009). Electronic tags also provide more detail regarding movement patterns than do conventional tags. These tags are mostly limited by battery duration, which can be between days to years.
Acoustic telemetry (e.g. Lucas and Baras 2000; Arnold and Dewar 2001;
Voegli et al. 2001) has been proven to be a useful tool in studying elasmobranchs
(Morrissey and Gruber 1993; Klimley et al. 2001; Heupel et al. 2004; Dewar et al.
2008), including P. microdon and G. garricki congeners, such as the speartooth shark
(Glyphis glyphis) (Pillans et al. 2009), dwarf sawfish (Pristis clavata) (Stevens et al.
2008), smalltooth sawfish (Pristis pectinata) (Simpfendorfer et al. 2010) and green sawfish (Pristis zijsron) (Peverell 2008). Acoustic telemetry involves internally or externally attached acoustic transmitters and corresponding acoustic receivers, which allow researcher to track animals in real-time for days, or monitor their movements
(i.e. recorded) for years. In acoustic monitoring, the movement of a tagged animal into the detection range of a receiver results in the recording of the time of its presence as well as the depth and temperature the transmitter was at, if environmental sensors are used. Detection range can vary from over 1000 m (Klimley et al. 2001) to a few metres, and is dependent upon transmitter strength (Klimley et al. 2001; Heupel
19 et al. 2006) and the surrounding environment (Voegeli et al. 1998; Clements et al.
2005; Lacroix et al. 2005; Heupel et al. 2006; Simpfendorfer et al. 2008a). As the emitted code is acoustic, environmental interferences can drown out, distort or absorb the signal. Useful in fresh and salt water, unlike radio transmitters, acoustic telemetry has been shown to be highly valuable in monitoring euryhaline species (e.g.
Simpfendorfer et al. 2005; Ng et al. 2007; Collins et al. 2007; Ortega et al. 2009) similar to P. microdon and G. garricki.
Unlike acoustic tags, satellite tags and most archival tags do not require a set array or to be manually tracked to relay data. This allows researchers to monitor animals across a much greater area, which is useful for long-ranging species.
Unfortunately, archival tags, which use recorded environmental data (e.g. light, temperature and depth) to estimate fish movements, can be unreliable in turbid waters or in times or locations when day lengths are equal (Teo et al. 2004). Alternatively, animals can be tracked remotely with satellite telemetry. Satellite (SAT) tags use satellites to aid in determining and relaying near-real time positions (Arnold and
Dewar 2001). Argos satellite-monitored tags calculate an animal’s approximate location with the assistance of low-orbiting polar satellites and radio transmissions.
The geolocation of a satellite tagged animal is calculated using the Doppler shift of four or more transmissions from a tag during a satellite’s pass (10-15 minute duration). Fewer transmissions can also produce a location estimate (Vincent et al.
2002). However, these estimates are less reliable (error of up to ~10º in diameter).
This tracking method is best used with animals that frequent surface waters, as the radio transmissions cannot penetrate far through the water column (Teo et al. 2004).
In aquatic environments, SAT tags work best with air breathing animals (e.g. Renaud and Carpenter 1994; Burns et al. 1998; Mate et al. 2006). However, SAT tags have
20 also proven to be effective with various elasmobranchs, including the great white shark (Carcharadon carcharius) (Bruce et al. 2006; Weng et al. 2007), basking shark
(Cetorhinus maximus) (Priede 1984) and salmon shark (Lamna ditropis) (Weng et al.
2005, 2008). Satellite tags are most appropriate for marine conditions, as SAT tags will commonly use conductivity switches to determine when to transmit. In a freshwater environment the conductivity switch is fooled into signaling the tag to transmit constantly, draining the battery and greatly shortening the life of the tag.
1.3 Pristis microdon and Glyphis garricki
Pristis microdon is one of seven members of the family Pristidae (sawfish) (currently under revision) (Faria 2007; Last and Stevens 2009), four of which have been confirmed to inhabit Australian waters (i.e. narrow sawfish [Anoxyprisits cuspidata],
Pristis clavata, P. microdon and Pristis zijsron) (Pogonoski et al. 2002; Thorburn et al. 2003, 2008; Peverell 2005; Last and Stevens 2009). Historically, P. microdon ranged throughout the Indo-West Pacific region; currently, viable populations are thought to be primarily confined to northern Australia (Pogonoski et al. 2002). Pristis microdon in Australia has been found to occur most commonly north of the 20˚ south latitude, from between the Pilbara region in Western Australia, north and east through to north-eastern Queensland (Last and Stevens 2009; Morgan et al. 2011). Pristis microdon is commonly found in inshore waters, including rivers, embayments, estuaries and tidal creeks (Compagno 1990; Peverell 2005; Thorburn et al. 2007).
Marine fisheries in the Northern Territory and Queensland have encountered sawfishes in their catches, although these captures consist mostly of A. cuspidata
(McAuley et al. 2005; Peverell 2005; Field et al. 2008). Until recently, P. microdon was thought to be an obligate freshwater fish (Last and Stevens 1994). However,
21 Peverell (2005, 2008), Thorburn (2006) and Thorburn et al. (2007) concluded that P. microdon inhabit rivers as juveniles and emigrate into marine environments just prior to, or during maturation.
Pristis microdon has been recorded up to 6600 mm total length (TL) and is estimated to attain at least 7000 mm TL (Compagno and Last 1998; Peverell 2008;
Last and Stevens 2009). Pristis microdon has a relatively evenly-spaced, broad (as an adult) and laterally tooth-lined rostrum, which protrudes anteriorly from its semi dorsal-ventrally flattened head (Compagno and Last 1998). This rostrum is used to sense, stun/kill and pin/manipulate prey (Wueringer et al. 2009). Pristis microdon has been documented to consume a variety of prey including ariid catfishes, mugiliids and crustaceans in Australia (Thorburn et al. 2007; Peverell 2008). Predation on adult
P. microdon is likely to be limited due to their comparatively large size and potentially damaging rostrum. However, juveniles have been observed to be taken by bull sharks (Carcharhinus leucas) (Thorburn 2006; Thorburn and Rowland 2008) and are suspected of being eaten by estuarine crocodiles (Crocodylus porosus) (Stirling
Peverell, pers. comm.).
Inhabiting similar waters as P. microdon, G. garricki has only been found to occur in the north-western region of Australia (from northern Western Australia, north and east to the eastern Northern Territory) (17˚21’S; 123˚27’E to ~11˚35’S;
~136˚25’E) and within the Fly River, Papa New Guinea (Thorburn and Morgan 2004;
Compagno et al. 2008; Pillans et al. 2009). Of the five species of the Glyphis
(Agassiz 1843) genus, only G. garricki and G. glyphis are known from Australian waters; the latter of which is known to inhabit Northern Territory and Queensland waters (Thorburn and Morgan 2004; Compagno et al. 2008; Pillans et al. 2009).
Although described as a river shark (Compagno 1984), having been reported in waters
22 with salinity as low as 2 ppt (Compagno et al. 2008; Pillans et al. 2009), G. garricki has been most commonly observed in tidally influenced and highly turbid estuarine and marine waters (Thorburn and Morgan 2004; Field et al. 2008, Pillans et al. 2009).
Glyphis garricki is a medium to large sized carcharhinid, reaching 2500 to
3000 mm TL. Glyphis garricki is distinguished from most other carcharhinids by a large second dorsal fin, which is one-half to three-fifths the size of the first, as well as by a longitudinal triangular (rather than crescentic) precaudal pit (Compagno and
Niem 1998; Thorburn 2006; Compagno et al. 2008). Reasonably large dorsal, pectoral and anal fins are hypothesised to aid in the manoeuvring of G. garricki in strong currents (Thorburn and Morgan 2004). A ‘waterline’ or ‘watermark’, the distinct transition boundary between dorsal and ventral colorations, is found on G. garricki and G. glyphis (Thorburn and Morgan 2004; Thorburn 2006; Compagno et al. 2008). Located immediately ventrally to the eye on G. glyphis and about an eyes diameter below the eye on G. garricki, this waterline is a useful characteristic for distinguishing these two species in the field. Eyes of G. garricki are unusually small
(Taniuchi et al. 1991; Thorburn and Morgan 2004; Thorburn 2006; Compagno et al.
2008) and are probably not relied on heavily for prey detection (Thorburn and
Morgan 2004). Thorburn and Morgan (2004) reported G. garricki to have a relatively large concentration of ampullae of Lorenzini compared to C. leucas. An increased number of ampullae would aid in prey detection in turbid waters or at night (Thorburn and Morgan 2004; Thorburn 2006).
Evidence of previous declines in numbers of P. microdon and the suspected small populations of G. garricki, has led to the recent listing of both species on international and Australian Federal and State legislation. Both species have been listed as Critically Endangered on the International Union for Conservation of Nature
23 and Natural Resources (IUCN) Red List of Threatened Species (Pogonoski and
Pollard 2003; Compagno et al. 2006). Pristis microdon has also been listed under
Appendix II of CITES, which prevents international trade of listed live animals and parts, with the exception of aquarium trade for conservation purposes (UNEP-WCMC
2010). Pristis microdon and G. garricki were also listed as Vulnerable and
Endangered, respectively, under Australia’s Environment Protection and Biodiversity
Conservation Act 1999 (the EPBC Act) (Department of Sustainability, Environment,
Water, Population and Communities 2011a, b). Listing of these species on the EPBC
Act allows for protection in Commonwealth waters, which begins three nautical miles offshore. Western Australia also added both species to the Fish Resources
Management Act 1994, under Schedule II (Full Protection), providing total protection in State waters to these fishes.
1.4 Status of and threats to Pristis microdon and Glyphis garricki
Sawfishes are believed to have declined in numbers globally; in recent decades becoming virtually extinct in several regions throughout their geographic range
(Pogonoski et al. 2002; Last and Stevens 2009). A decline in the Australian P. microdon population is evident from both a retraction in home ranges as well as a decline in catch rates (Pogonoski et al. 2002; Stevens et al. 2005). Historical population levels of G. garricki are unknown as the species was only discovered in
Australia in 1986 (Taniuchi et al. 1991), and in Western Australia in 2002 (Morgan et al. 2004; Thorburn and Morgan 2004). Glyphis garricki populations are believed to be small (Pogonoski et al. 2002; Thorburn and Morgan 2004), which may be a sign of a previous decline in numbers. However, due to a paucity of information and captures of G. garricki (only 32 G. garricki were recorded between 1986 and 2006 [Pillans et
24 al. 2009]), it is not possible to ascertain population trends (Compagno 1984;
Pogonoski et al. 2002; Thorburn and Morgan 2004; Pillans et al. 2009). Field et al.
(2008) (using data from an onboard fishery observer program in the Northern
Territory) suggested that current Australian Glyphis spp. numbers may not be as low as currently thought. Misidentification (Compagno 1984; Pogonoski and Pollard
2003; Field et al. 2008) as well as minimal sampling effort and studies targeting
Glyphis spp. (e.g. Taniuchi et al. 1991; Thorburn et al. 2003; Thorburn and Morgan
2004; Martin 2005; Thorburn 2006; Field et al. 2008) may have resulted in the underestimation of Australian Glyphis spp. population sizes. Indeed, further research is required to substantiate current distribution and population estimates of G. garricki and P. microdon.
Threats likely responsible for the decline of P. microdon and G. garricki include fishing pressure (for at least P. microdon) and habitat degradation. Pristis microdon is a demersal batoid with a tooth lined-rostrum and thus is highly susceptible to certain fishing gears. Pristis microdon has been captured in both prawn trawls and gillnet-fisheries (McAuley et al. 2005; Peverell 2005; Stevens et al. 2005;
Field et al. 2008). Recreational and traditional fishers have also been documented to catch this species, as P. microdon readily accepts baited hooks and are culturally significant to traditional cultures (Thorburn et al. 2003; Morgan et al. 2004). Stevens et al. (2005) reported barramundi gillnet fisheries to have the highest percentage of total Pristis by-catch (80.2 %), followed by trawling (16.6 %), line (9.2 %) and recreational gears (0.3 %).
Habitat degradation is known to greatly impact nearshore euryhaline species such as P. microdon and G. garricki (Compagno and Cook 1995). Pollution and deforestation can greatly affect elasmobranchs, and has caused major declines and
25 extinctions of P. microdon and G. garricki congeners in countries such as Thailand and Papa New Guinea (Compagno et al. 2006). Seasonal flooding in Australia has been observed to transport high concentrations of copper sulfate from a mine into the
Leichardt River, Queensland, a known P. microdon habitat. Although no deceased elasmobranchs were observed during this spill, concentrations within the river during this period were measured at lethal levels (Stirling Peverell, per comm.). Mining, water diversion/abstraction and hydroelectric damming projects throughout northern
Australia have impacted rivers inhabited by P. microdon. Damming of the Ord River in Western Australia caused the seasonally flowing river to flow perennially, decreased the intensity of flooding events and now restricts movement of migratory fishes (Wolanski et al. 2001; Doupé and Pettit 2002). Such transformations are known to disrupt fish migrations, alter species survivorship and behaviour as well as change the species dynamics of the affected area (Banks 1969; Raymond 1979;
Drinkwater and Frank 1994; Gehrke et al. 2002). Unfortunately, it is uncertain how changes to the habitats of P. microdon or G. garricki affect these species, as little is known about their habitat use. While areas in the western Kimberley, such as the
Fitzroy River and King Sound are relatively pristine, they are under ongoing threat of damming, mining and water abstraction programs, such as those discussed by Lindsay and Commander (2006) and CSIRO (2009).
1.5 Aims of the study
There is an urgent need to understand and obtain fine-scale and long-term data on the movements and habitat use of the critically endangered P. microdon and G. garricki, as these species continue to face the threat of extinction. Although, Phillips et al.
(2011) reported the genetic diversity of the Fitzroy River P. microdon population to
26 appear relatively high (compared to other endangered elasmobranchs), a better understanding of the ecology of this species is needed to increase the efficacy of species management to maintain or increase the health of this population. To increase the effectiveness of land and fishery management strategies for the conservation of P. microdon and G. garricki, this study proposes to achieve the following aims:
Tidal action, precipitation and air temperature are three predominant
environmental variables that greatly fluctuate throughout the year in
northern Australia. These variables are hypothesised to be the driving
forces of the chemical and physical properties of the aquatic environment
in the area. As a detailed understanding of such processes is important for
habitat use studies, this thesis aims to test this hypothesis by documenting
the temporal and spatial patterns of numerous environmental variables
related to the aquatic environment, and explore their interrelationships
(Chapter 2).
Recent evidence has suggested global populations of P. microdon and G.
garricki are declining and/or are small in number, respectively. As such,
it is hypothesised that the Fitzroy River and King Sound populations are
declining and/or are small in number. It is the aim of this study to
document the trends of these populations to confirm this hypothesis.
(Chapter 3).
Tracking the movements of P. microdon and G. garricki can be difficult,
due to the heterogeneity and remoteness of their home range and the
27 limits of currently available tagging methods. However, it is hypothesised
that mark and recapture methods, passive acoustic monitoring and satellite
telemetry will complement one another and be effective methods for
monitoring P. microdon and G. garricki throughout their home range.
This study aims to test this hypothesis and determine the utility and
efficacy of these tagging methods on these species in the Fitzroy River
and King Sound (Chapters 4-5).
The effects of habitat modification on P. microdon and G. garricki is
unclear, but are hypothesised to influence the habitat use behaviours of
these species. This study aims to use environmental, catch and tagging
data to document the habitat use patterns of P. microdon and G. garricki,
and determine how they are influenced by environmental changes
(Chapters 2-5).
28 Chapter 2: Environments of the Fitzroy River and
King Sound, Western Australia
2.1 Introduction
Detailed knowledge of a study area is a necessary component in behavioural ecology field studies. Identification of the physical components and processes of a habitat is of great importance and relevance when explaining habitat use, habitat selection and movement patterns of elasmobranchs (Morrissey and Gruber 1993; Ackerman et al.
2000; Heupel et al. 2003; Sims et al. 2006; Chalfoun and Martin 2007; Collins et al.
2007; Dewar et al. 2008; Kerford et al. 2008). In these studies it is important to focus on the spatial and temporal heterogeneity of an area (e.g. Fritz et al. 2003; Sims et al.
2006; Heithaus et al. 2006; Heithaus and Dill 2006; Chalfoun and Martin 2007). This is because the patchiness of an environment can dictate the behaviour of a fish, as fish will often respond to fluctuating resources to minimise stress and maximise resources, either by moving to a more optimal habitat or by altering their habitat use behaviour
(Imbrock et al. 1996; Matern et al. 2000; Hardiman et al. 2004; Dewar et al. 2008;
Heupel and Simpfendorfer 2008). However, it is imperative that these studies identify the spatial and temporal scales appropriate for the hypothesis in question, as observed behavioural patterns can change depending on the scale of the study (Fritz et al. 2003;
Heithaus and Dill 2006; Chalfoun and Martin 2007). Incorporating multiple scales into a study can be an effective means to capture the most meaningful results (García-
Charton et al. 2004). Incorporation of multiple scales in analyses of habitats and home ranges is especially important in investigations of long-lived euryhaline species, such as the freshwater sawfish (Pristis microdon) (Thorburn et al. 2007; Last and
29 Stevens 2009) or the northern river shark (Glyphis garricki) (Thorburn and Morgan
2004; Last and Stevens 2009), which inhabit constantly changing environments, like northern Western Australia (Semeniuk 1980, 1981; Taylor 1999; Wolanski and
Spagnol 2003; Lindsay and Commander 2006).
The Fitzroy River arises in the King Leopold Ranges in the north-western
Kimberley, Western Australia, and meanders for some 600 km before flowing into
King Sound (1733’12”S, 12335’20”E), south of Derby (Fig. 2.1). Numerous tributaries and anabranches including the Margaret, Mitchell, Little Fitzroy and Hann
Rivers, as well as Christmas Creek and Snake Creek deviate from the main channel into the surrounding environment. The environment surrounding the Fitzroy River changes considerably with movement downstream from the headwaters, transforming from the rocky steppes of its origin, to sparse and sand-rich areas downstream of
Fitzroy Crossing, to the lush pools downstream of the Camballin Barrage (Storey et al. 2001). The most notable changes occur with approach to the tidally influenced estuarine pools. Moving downstream into the estuarine region of the river (lower 17.5 km of the river), the breadth of the river greatly increases from between 0.05 and 0.08 km in the freshwater pools to 0.1 to 0.6 km in the estuarine pools, depending upon season and tidal action. Additionally, the substrate transforms into a more fine-grade composition (increased ratio of silt and mud in place of sand-rich and relatively course substrate). The vegetation also changes from a comparatively lush woodland composed of Eucalyptus spp., Ficus spp. and Acacia spp. trees (Storey et al. 2001), to a more sparse woodland dominated by Melaleuca spp. with intermittent Eucalyptus spp. trees. Only minor road crossings and the Camballin Barrage, a weir located 160 km upstream of the river mouth, obstruct the path of the relatively pristine river. The
Camballin Barrage is relatively small in size but has been suggested to disrupt up and
30 downstream movement of fishes for over nine months of the year (Morgan et al.
2005). Snake Creek is an anabranch of the river, which flows from the pool upstream of the barrage to downstream of Myroodah Crossing (see below). Snake Creek is the only current diversion for the barrage.
The Fitzroy River is one of Australia’s largest rivers and Western Australia’s longest, with a catchment basin covering over 93,800 km2 (The Australian
Government Department of the Environment, Water, Heritage and the Arts 2009).
Discharge of the Fitzroy River can be extremely variable. Annual discharge ranged from 819 (2005) to 20,400 GL yr-1 (2000) between 2000 and 2009 (Hydstra database- time-series data 2009). During the wet season, which typically occurs between
December and April, river stage height in the freshwater regions can reach over 10 m above its dry season height, and the river often spills onto its wide flood-banks
(Taylor 1999). However, the wider breadth and shorter banks of the river in the lower estuarine pools (~1-3 m, rather than ~6-10 m as observed upstream) does not allow for such a large depth increase during the wet season. In the dry season, which typically occurs between May and November, discharge and stage height greatly decreases. In the winter months (encompassed by the dry season), air temperature is at its coldest. However, being a tropical environment, the lower region of the river
(i.e. Derby) still obtains a mean monthly maximum temperature of over 30°C
(Commonwealth of Australia Bureau of Meteorology 2010). Air temperatures and evaporation rates increase with progression of the dry season (evaporation rates are lowest in June and greatest in December), resulting in a decrease in river flow and depth. These changes transform the flowing river into a series of pools connected by shallow runs, commonly <0.5 m in depth, in the dry season. Pools often become completely isolated from one another during the mid to late dry season. River flow
31 typically ceases in the late dry season and evaporating waters are then only replenished through the local alluvial aquifer (Lindsay and Commander 2005).
Along with the Fitzroy River, the Meda, May, Robinson and the smaller
Fraser Rivers as well as hundreds of tidal creeks, dot the coastline of King Sound.
King Sound is fed by the Indian Ocean and is mostly marine. Salinity in the sound is likely to change during the wet season when substantial increases in run-off and river discharge penetrate the sound. Approximately 100 km long and over 50 km wide (at its widest point), King Sound is a large embayment that is dominated by >11-m mixed-macro tides. King Sound is a relatively shallow water body, with a mean depth of 18 m (Wolanski and Spagnol 2003). The majority of the upper (i.e. southern) King
Sound (extends approximately from the Fraser River in the west, east to Stokes Bay) and its tidal creeks are drained during extreme low tides. This leaves only a few navigable channels amongst a sea of sandbars. A deeper channel, measuring up to 50 m deep (Wolanski and Spagnol 2003), extends inward from the narrow mouth of the sound through to the Point Torment area (see Fig. 2.1). The upper sound, which encompasses the mouths of all major rivers in the area, typically consists of mud and salt flats that are covered by dense mangrove forests (Fig. 2.1) (Semeniuk 1980, 1981,
1983). Salt flats, which border the majority of the larger tributaries, maintain an elevated salinity of groundwater that is often toxic, even to local mangrove species
(Semeniuk 1983). The sand-flats tend to be more prevalent outside of the creek mouths in King Sound (Wolanski and Spagnol 2003). High turbidity levels in the upper King Sound form a distinctive front with the less turbid, blue-hued and deeper waters of the sound. This turbidity maximum has been suggested to have originated as a result of the composition of the flocculants being suspended, and is maintained by a combination of tidal action and wind generated waves (Wolanski and Spagnol
32 2003). Northern areas of the sound, which are in close proximity to the mouth of the sound, are characteristically less turbid and coloured with a milky-blue hue.
Numerous studies have focused on the Fitzroy River and King Sound within the last 30+ years, encompassing investigations into the flora (e.g. Semeniuk 1980,
1981, 1983; Storey et al. 2001), fauna (e.g. Webb and Zhang 1997; Storey et al. 2001;
Morgan et al. 2002, 2004, 2005, 2008; Thorburn et al. 2004, 2007, 2008; Thorburn and Morgan 2004; Thorburn 2006) and geology/hydrology (e.g. Jennings and
Coventry 1973; Jennings 1975; Wyrwoll 1992; Taylor 1999; Wolanski and Spagnol
2003; Lindsay and Commander 2005). Alongside these studies, the Department of
Water, Government of Western Australia has monitored and provided invaluable records on river flow, temperature and water chemistry in the Fitzroy River since before 1960. Unfortunately, to the author’s knowledge, no published study has yet commented on multi-scale temporal and spatial changes influencing the physicochemical components of the Fitzroy River and King Sound. Information regarding the estuarine region of the Fitzroy River is particularly lacking.
Fishes inhabiting the Fitzroy River and King Sound are likely to change their habitat use patterns in conjunction with the extreme changes in environmental variables that occur in this area, as a form of behavioural homeostasis to maximise fitness. To better understand observed changes in habitat use of fishes within the
Fitzroy River and/or King Sound, this study aimed to document temporal and spatial changes of environmental variables that may influence aquatic life. It is reasonable to assume that environmental variables within the river and sound are likely to experience biannual changes due to the annual occurrence of the wet and dry seasons.
However, the natural cycles of other prominent factors (referred to here as ‘primary variables’) such as tides (Meade 1966; Wolanski and Spagnol 2003) and air
33 temperature (Pilgrim et al. 1998; Langan et al. 2001; Malcolm et al. 2004; Webb and
Nobilis 2007), which also influence aquatic environmental variables (referred to here as ‘secondary variables’), do not follow a wet/dry temporal pattern. As such, secondary variables (i.e. water temperature, salinity, turbidity, pH and DO) may not take on a biannual pattern, but are likely to follow temporal patters similar to those of the primary variables that are the most influential. However, it was hypothesised that the strength of the relationship between primary and secondary variables may be altered by temporal changes in the prominence of the primary variables. Additionally, it was also hypothesised that changes in the strength of the relationships between primary and secondary variables would also vary depending upon the region in which they occur (the term region is used in this thesis as a reference to the macrohabitats of the study area, namely King Sound and the estuarine as well as freshwater pools of the Fitzroy River), due to the physical variations (i.e. substrate composition, hydrological geometry, vegetated cover) of each region. To test these hypotheses, this study documented numerous environmental variables within the Fitzroy River and King Sound in 2007 through 2009. Each variable was explored and discussed in detail to aid in future biological studies. Outcomes from this study will also be of importance in other chapters of this thesis, which investigated the habitat use of P. microdon and G. garricki in the western Kimberley.
2.2 Materials and Methods
2.2.1 Study Sites
Water temperature, salinity, turbidity, light intensity, pH and dissolved oxygen (DO) were monitored in eight pools within the lower 160 km of the Fitzroy River, between
34 the river mouth and the Camballin Barrage in 2007 through 2009 (Fig. 2.1a). Notes on the vegetation and substrate were also recorded. As no defined mouth of the river exists, for the purposes of this study, the Fitzroy River ‘mouth’ was set at Lower
Pelican Pool (name assigned by the author and given in respect to the neighboring upstream pool, Pelican Pool) (17°21'47"S; 123°20'55"E) (see Fig. 2.1a). The last major anabranch of the river, the Cuttings, rejoins the main channel at the designated river mouth. The Fitzroy River ‘mouth area’ was considered the general area that encompasses the estuarine waters between the start of the river delta (4 km upstream of the river mouth) to the end of the river plume (extends tens of kilometres into King
Sound). Five of the monitored pools (Lower Pelican Pool, i.e. the river mouth; Milli
Milli, 5 km upstream of the river mouth; Snag Pool, 6.2 km upstream of the river mouth; Telegraph Pool, 11.6 km upstream of the river mouth; Langi Crossing, 14.9 km upstream of the river mouth) were located in the tidally influenced lower 17.5 km of the river (Fig. 2.1a, b). Other monitored pools were located in the freshwater (<1.0 ppt) reaches of the river (Myroodah Crossing, 123 km upstream of the river mouth;
Camballin Pool 156.5 km upstream of the river mouth, name assigned by author) (Fig.
2.1a, c).
Water temperature, salinity, turbidity, DO and depth were measured at up to
35 sites throughout King Sound in 2007 through 2009 (Fig. 2.2). Notes on substrate composition were also recorded. Environmental measurements were taken opportunistically during fishing trips, which occurred in October 2007 in the Point
Torment, Stokes Bay and North-East (NE) King Sound areas, July 2008 in the Point
Torment, South-East (SE) and South-West (SW) King Sound areas and October 2009 in the NE King Sound area (Fig. 2.2). Sites sampled within King Sound included macro-tidal estuarine creeks and nearshore channels. A single offshore site (>5 km
35 from the shore) within the Point Torment area was also sampled (Fig. 2.2). Sampled
sites were divided into five areas (listed above) based on their location and habitat
types (Fig. 2.2).
King Sound b
Western Australia
Lwr Pelican Pl 17°30'S Snag Pl Milli Milli
Telegraph Pl Langi Crs. Willare St.
Fitzroy River
Myroodah Crs. 1 18°0'S Camballin Pl
¸ Looma St. 0 5 10 20 30 40
Kilometers Camballin Barrage St.
123°30'E 124°0'E 124°30'E
Fig. 2.1. (a) Map of sampled sites in the Fitzroy River (2007 to 2009), and satellite photos of (b)
estuarine and (c) freshwater pools in the dry season at similar magnification.
36 16°30'S
Robinson River
King Sound 17°0'S
Meda River
May River Fraser River
b 17°30'S Western Australia Fitzroy River ¸ 0 5 10 20 30 4040
Kilometers 123°0'E 123°30'E 124°0'E
Fig. 2.2. Sampled sites (red dots) and areas (shaded) within King Sound (2007 to 2009). Green
shading = NW King Sound area, orange shading = SW King Sound area, yellow shading = SE King
Sound area, red shading = Point Torment area and blue shading = Stokes Bay area.
37
Plate 1. (a) Myroodah Crossing Pool and run and (b) Snag Pool in the dry season; (c) high and (d) low tide in a King Sound creek; e) Camballin Barrage; f) Myroodah Crossing.
38 2.2.2 Depth
River pool depths were documented at each sampled site within the Fitzroy River in
June 2007, with exception of Telegraph Pool (Telegraph Pool was incorporated into the study at a later date), using a water depth gauge. Depth was recorded at intervals of 0.05 km during pool transects, which were run 0.5 to 3 km lengthwise (depending upon the length and depth of the pool) through the middle of each pool. Transect pathways typically followed that of the main channel in each pool. Pools’ depths derived from transects should only be treated as approximations, as measurements were typically taken once.
River stage height and discharge were collected on a continuous basis for the length of the study and were supplied by the Department of Water, Government of
Western Australia. River stage height was measured at the Camballin Barrage station
(referred to in the data as the ‘Fitzroy Barrage’) (see Fig. 2.1a). River discharge data was collected from the Willare station (see Fig. 2.1a), as it was not available for the
Camballin Barrage station. Water depths were also measured opportunistically in
King Sound via an echo sounder during fishing events.
2.2.3 Temperature
Pool temperatures were continuously logged at three-hour intervals in Milli Milli and
Myroodah Crossing between 2007 and 2009, and in Snag Pool, Langi Crossing and
Camballin Pool between 2007 and November 2008, by HOBO Water Temp Pro V2
Data Loggers (Onset Computer Corporation, Bourne, MA, USA). Loggers were placed on moorings at approximately one to two metres depth in all monitored pools except Telegraph Pool (Telegraph Pool was incorporated into the study at a later date). In addition, HOBO Pendant Loggers (Onset Computer Corporation, Bourne,
39 MA, USA) were deployed to continuously log temperature through the water column at 0.5-m intervals. Pendant Loggers were deployed in Snag Pool, Langi Crossing,
Myroodah Crossing and Camballin Pool (see Fig. 2.1a) in June and October 2008.
Pendant loggers recorded in 15-min intervals for one to three days, and again in
Myroodah Crossing and Telegraph Pool, recording in one-hour intervals in late
November 2008 to April 2009.
Surface and bottom water temperatures were measured opportunistically using an YSI-30 temperature and salinity probe (Yellow Springs, Ohio, USA) in the Fitzroy
River and King Sound (Fig. 2.1a, 2.2; see Chapter 3). Measurements were typically recorded in June/July and October/November of 2007 to 2009. Surface and bottom water temperatures were also recorded in June 2007, during pool transects, which extended 0.5 to 3.0 km lengthwise through the middle of all monitored pools, excluding Telegraph Pool (Telegraph Pool was incorporated into the study at a later date). Water temperature was measured at intervals of 0.05 km during transects. Air temperature was measured daily between 2007 and 2009, at the Derby Aero station
(17º22’12” S, 123º39’36” E). Air temperature data was made available through the
Bureau of Meteorology, Commonwealth of Australia (2010).
2.2.4 Salinity
Salinity was measured opportunistically in surface and bottom waters of each sampled site in the Fitzroy River and King Sound using an YSI-30 temperature and salinity probe (Yellow Springs, Ohio, USA). Measurements were typically taken in June/July and October/November of 2007 to 2009. The salinity of surface and bottom waters were also measured in June 2007, during pool transects, which extended 0.5 to 3.0 km lengthwise through the middle of each monitored pool, excluding Telegraph Pool
40 (Telegraph Pool was incorporated into the study at a later date). Salinity was measured at intervals of 0.05 km during transects.
2.2.5 Turbidity and light
Turbidity and/or visibility were measured opportunistically in June/July and
October/November at each monitored pool, except Lower Pelican Pool and Milli
Milli, using a Secchi Disc in 2007, and an Oakton TN-100/T-100 Portable
Turbidimeter (Oakton, Vernon Hills, IL, USA) in 2008 and 2009. Turbidity was also measured in the SW King Sound area in July 2008, and in the NW King Sound area in
September 2009.
Light intensity was continuously logged via HOBO Pendant Loggers (Onset
Computer Corporation, Bourne, MA, USA) at multiple depths set 0.5 m apart.
Pendant Loggers were deployed in Snag Pool, Langi Crossing, Myroodah Crossing and Camballin Pool (see Fig. 2.1a) in June and October 2008. Pendant loggers recorded in 15-min intervals for one to three days, and again in Myroodah Crossing and Telegraph Pool, recording in one-hour intervals in late November 2008 to April
2009. During deployments in June and October 2008, a logger was placed 0.3 m above the surface to aid in normalising light intensity measured at different times.
Sensitivity of light sensors varied between wavelengths of ~200 and 1200 nm, but peaked at 900 nm.
2.2.6 Dissolved oxygen and pH
Dissolved oxygen was measured using an Oakton PCD650 meter (Oakton, Vernon
Hills, IL, USA). Dissolved oxygen was measured in all pools except Telegraph Pool in June 2007 (Telegraph Pool was incorporated into the study at a later date), and
41 again in June 2008, with exception of Lower Pelican Pool, due to inaccessibility.
Dissolved oxygen was measured again in November 2008, in all pools excluding
Milli Milli and Lower Pelican Pool, due to their inaccessibility. Dissolved oxygen levels were measured in the SW King Sound area in July 2008.
Surface water pH was measured in June and November 2008, in all pools excluding Lower Pelican Pool and Milli Milli, using an Oakton PCD650 meter
(Oakton, Vernon Hills, IL, USA). Analysed pH data was supplemented with measurements supplied by the Department of Water, Government of Western
Australia, due to the small number of samples taken in the freshwater pools. These samples were taken at the Camballin Barrage and the Looma station (named by the
Department of Water, Government of Western Australia; located 5 km downstream of
Myroodah Crossing) (see Fig. 2.1a).
2.2.7 Tidal cycle
Tide data for the Derby Jetty was supplied by the Department of Planning and
Infrastructure, Government of Western Australia for the study period. Real-time visual observations, made throughout the dry season in 2007 to 2009, were used to estimate the time difference of high tide between the Derby Jetty and the estuarine pools of the Fitzroy River. For this study, the start of an incoming tide was defined as the point when tidal flow overpowered river flow. The time of the incoming tide was noted when nearshore sediment plumes reversed in flow direction. High tide was noted when sediment plume direction was unilateral.
42 2.2.8 Data analysis
Statistical analyses were conducted using SigmaPlot for Windows Version 11.0
(Systat Software, Inc., Chicago, IL, USA). Years were pooled for analyses to increase the power of these tests. Means are reported with standard errors in the format: mean ± se, unless otherwise stated. The strength of the relationships between water temperature and river flow, air temperature and tides were analysed using
Spearman’s rank correlation, as water temperature on an annual scale had a bimodal distribution. These relationships were reanalysed for each season (i.e. early dry, late dry, early wet, late wet) using a Pearson product moment correlation (water temperature was normally distributed for seasonally specific data). To explore the influence of temporal scale on these relationships, the data analysed in these correlations consisted of daily, weekly, monthly and/or seasonal means. Each analysis was run for all monitored pools on a regional basis (i.e. estuarine and freshwater pools) and for the river as a whole to explore the influence of spatial scales on these relationships. Differences in daily water temperature variations between pools and seasons, and differences in water temperature, salinity and oxygen between locations in King Sound and its tributaries were investigated using a Kruskal-Wallis one-way ANOVA with a Dunn’s post hoc test. Comparisons between surface and bottom water temperature, salinity and DO were examined with the use of a paired
Student’s t-test or Wilcoxon signed rank test, when appropriate (Student 1908,
Conover and Iman 1981). Significant differences in environmental variables between locations and seasons were analysed using a grouped Student’s t-test or a Mann-
Whitney rank sum test, when appropriate (Student 1908; Conover and Iman 1981).
As many of the statistical tests were non-parametric, an increased level of Type II errors were likely to occur and this should be considered when interpreting the results.
43 2.3 Results
2.3.1 Depth
Water level (river stage height), river discharge, as well as the onset and duration of the wet season varied annually (Fig. 2.3-2.5). Mean wet season stage height was large in 2002, 2004, 2006 and 2008 (13.2 ± 0.12 to 14.8 ± 0.10) in comparison to 2003,
2005 and 2007 (11.1 ± 0.04 to 12.6 ± 0.11). The occurrences of peaks in river depth and discharge varied between years but were constrained between December and
April (Fig. 2.3, 2.4). Annual river discharge (December to November) was <5.0 x 106
ML in 2003, 2005, 2007 and 2008, and >7.0 x 106 ML in 2002, 2004, 2006 and 2009
(Fig. 2.5).
Depths of pools were found to increase with distance from the mouth in the estuarine region. Maximum depths of the estuarine pools ranged between 2.0 (Lower
Pelican Pool) and 3.4 m (Langi Crossing) in June 2007 (Fig. 2.6a). Myroodah
Crossing and Camballin Pool were observed to be deeper than most estuarine pools, with measured depths of up to 3.5 (Camballin Pool) and 4.2 m (Myroodah Crossing) in June 2007 (Fig. 2.6b). Additionally, small (few metres in length) but deep pockets or holes were found along the banks. These holes were up to 3 m in depth in the estuarine pools and up to 6 m in depth in the freshwater pools.
44 20 2002 2006 18 16 14 12 10 20 2003 2007 18 16 14 12 10 20 2004 2008 18
Stage height (m) height Stage 16 14 12 10 20 2005 2009 18 16 14 12 10 Dec Jan Feb Mar Apr May Dec Jan Feb Mar Apr May Date Date Fig. 2.3. Maximum daily stage height (m) at Camballin Barrage station, Fitzroy River for the 2002 to 2009 wet seasons; data courtesy of the Department of Water, Government of Western Australia.
12000 6000 2002 2006 8000 4000 4000 2000 0 0 6000 ) 2003 2007 -1 4000
s
3 2000 0 6000 2004 2008 4000
Discharge(m 2000 0 6000 2005 2009 4000 2000 0 Dec Jan Feb Mar Apr May Dec Jan Feb Mar Apr Date Date
Fig. 2.4. Maximum daily discharge (m3 s-1) at Willare station, Fitzroy River for the 2002 to 2009 wet seasons; data courtesy of the Department of Water, Government of Western Australia.
45 12000
10000
8000
6000
4000
River discharge (GL) discharge River
2000
0
2002 2003 2004 2005 2006 2007 2008 2009 Year
Fig. 2.5. Total annual (December to November) river discharge (GL) at Willare station, Fitzroy
River for 2002 to 2009; data courtesy of the Department of Water, Government of Western Australia.
46 0
1
2
Depth (m) Depth
3 Lower Pelican Pool Milli Milli Snag Pool Langi Crossing 4 -600 -400 -200 0 200 400 600
0
1
2
Depth (m) Depth 3
4
Myroodah Pool Camballin Pool 5 -400 -200 0 200 400 600 Distance from mooring (m)
Fig. 2.6. Depth profiles of the (a) estuarine and (b) freshwater pools. Depth profiles extend to the maximum detection range up (+) and downstream (-) of the moored acoustic receivers (see Chapter 4)
(i.e. the deepest point in a pool).
47 2.3.2 Temperature
Water temperature in the Fitzroy River ranged between 15 (July) and 38.1C
(November and December) during the study period (Table 2.1, Fig. 2.7). The greatest changes in mean weekly river temperature for each sampled year occurred roughly in
April through May (referred to as ‘late wet’) and in August through November
(referred to as ‘late dry’). Between these periods weekly mean (± s.d.) river temperature of monitored pools remained relatively constant between 20.4C (± 1.4) to 22.1C (± 2.5) and 31.6C (±1.4) to 31.8C (±1.2) in the winter (referred to as
‘early dry’) and summer (referred to as ‘early wet’) months, respectively (Fig. 2.7).
Mean daily temperatures were not significantly different between Snag Pool and Milli
Milli (p = 0.251) or between Myroodah Crossing, Camballin Pool and Langi Crossing
(p > 0.105). However, mean daily temperatures were significantly different between
Milli Milli and the freshwater pools (i.e. Myroodah Crossing, Camballin Pool) as well as Langi Crossing (p < 0.022), and between Snag Pool and Langi Crossing (p =
0.026). Interestingly, no significant difference was observed between Snag Pool and
Myroodah Crossing or Camballin Pool (p > 0.262).
Table 2.1. Minimum and maximum water temperatures of monitored pools in 2007 to 2009
Site Minimum Maximum Milli Milli 15.0 38.1 Snag Pool 15.3 36.8 Langi Crossing 17.4 34.5 Myroodah Crossing 18.1 34.5 Camballin Pool 17.8 35.5
48 40
35
30
C)
o
25
Temperature ( Temperature 20
Milli Milli 15 Myroodah Crossing Air King Sound 10
Jun-07 Oct-07 Feb-08 Jun-08 Oct-08 Feb-09 Jun-09 Oct-09 Date
Fig. 2.7. Recorded King Sound, Fitzroy River (Milli Milli = estuarine representative, Myroodah
Crossing = freshwater representative) and air temperatures between 2007 and 2009; air temperature was courtesy of the Bureau of Meteorology, Commonwealth of Australia (2010).
Logged daily temperature fluctuations decreased with distance from the river mouth (Fig. 2.8). Fluctuations in daily temperature increased significantly between the late wet and late dry months (p < 0.05) for all monitored pools except Langi
Crossing (Fig. 2.8). The greatest differences in daily temperatures were 8.5, 7.3 and
5.4 °C in Milli Milli, Snag Pool and Telegraph Pool, respectively. Maximum differences in daily temperatures were only 3.6, 2.6 and 3.3 °C in Langi Crossing,
Myroodah Crossing and Camballin Pool, respectively. The range of daily water temperature in the late dry season was found to decrease in periods of large tides
(>9.5 m in Milli Milli) in the lower estuarine pools, a result of decreased maxima temperatures (Fig. 2.9). This decrease in daily water temperature occurred one to two days after tidal waters appeared (Fig. 2.9). The influence of tides on water temperature was negligible during other seasons.
49 Thermal stratification increased between the early and late dry season in the freshwater pools and in Langi Crossing. The difference between surface and bottom water temperatures (dT) was not significant in Myroodah Crossing (p = 1, d.f. = 16,
-1 Z = 3.1 m ± 0.2) and Langi Crossing (p = 0.146, dT = 0.02C m ± 0.18, n = 28), and was only borderline significant in Camballin Pool (p = 0.031, d.f. = 16, dT = 0.03C m-1 ± 0.01) in June 2007. Comparatively, this difference was found to be highly significant in all three pools (Myroodah Crossing: p < 0.001, d.f. = 16, dT = 0.1C m-1
± 0.04; Camballin Pool: p < 0.001, d.f. = 9, dT = 0.5C m-1 ± 0.05; Langi Crossing: p
= 0.002, n = 17, dT = 0.24C m-1 ± 0.16) in October and November 2007. Measured differences between bottom and surface water of Lower Pelican, Milli Milli and Snag
Pool varied between -0.5 and 0.3C m-1 in the dry season. Differences between surface and bottom water temperatures in these pools were not statistically analysed, due to small sample sizes (n = 6 to 15).
6 Milli Milli Snag Pool Telegraph Pool
C) 5 Langi Crossing
o Myroodah Crossing Camballin Pool 4
3
2
1
Mean diel temperature range ( range temperature diel Mean
0
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Month
Fig. 2.8. Monthly mean (± s.e) diel temperature ranges of monitored Fitzroy River pools.
50 12.0 36 Tide Water temperature
11.5 34
C)
o 11.0 32
10.5 30
Tidal height (m) height Tidal 10.0 28
Water temperature ( temperature Water
9.5 26
9.0 24
06-Oct-08 13-Oct-08 20-Oct-08 27-Oct-08
Date
Fig. 2.9. Tide and water temperature (ºC) profiles in Milli Milli in the late dry season 2008.
Water temperature in the river was found to be associated with air temperature, river discharge and maximum tidal height, but at varying strengths and scales. Air temperature followed a similar patter to water temperature and ranged between 18 and 35.3C during the study period (Fig. 2.7). Accordingly, air temperature was found to be strongly correlated to river temperature at nearly all temporal and spatial scales (Table 2.2). The strength of the correlation between air and water temperature was observed to vary seasonally and was weakest during the early wet season. Strength of the relationship between air and water temperature was weaker at a daily scale in comparison to a weekly scale (Table 2.3). Similarly, the relationship between river discharge and water temperature changed in strength and polarity with seasons and was only strongly correlated during the late wet season
(Table 2.3). Water temperature appeared to be mostly independent of maximum tidal height. A consistent, minor and negative correlation between these two variables was observed to only develop in the late dry season in the estuarine pools at a daily scale
(Table 2.3).
51 Table 2.2. Correlation coefficients (p < 0.05, score not given if p > 0.05) of river water and air
temperatures, as well as river discharge of monitored pools at varying temporal and spatial
scales
Variable Site Daily Weekly Monthly Seasonal Milli Milli 0.86 0.88 0.86 0.92 Snag Pool 0.86 0.89 0.89 - Langi Crossing 0.88 0.91 0.91 0.89 Myroodah Crossing 0.9 0.94 0.94 0.92 Air temperature Camballin Pool 0.82 0.86 0.89 0.81 Estuarine pools 0.86 0.89 0.87 0.89 Freshwater pools 0.84 0.92 0.9 0.88 River 0.88 0.91 0.93 0.89 Milli Milli 0.36 0.39 0.57 - Snag Pool 0.2 - - - Langi Crossing 0.14 - - - Myroodah Crossing 0.11 - - - River discharge Camballin Pool 0.2 0.25 - - Estuarine Pools 0.21 - - - Freshwater Pools 0.16 - - - River 0.14 - - -
52 Table 2.3. Seasonally separated correlation coefficients (p < 0.05, score not given if p > 0.05) of
river water and air temperature, river discharge and tidal height at varying spatial scales
employing daily and weekly means
EW, early wet; LW, late wet; ED, early dry; LD, late dry; Estuarine pools, mean of Snag Pool and
Milli Milli; Freshwater pools, mean of Myroodah Crossing and Camballin Pool; River, mean of all
monitored pools.
Daily Weekly Variable Site EW LW ED LD EW LW ED LD Milli Milli 0.52 0.85 0.72 0.80 0.60 0.88 0.84 0.90 Snag Pool 0.45 0.86 0.62 0.76 0.75 0.90 0.79 0.91 Langi Crossing 0.44 0.87 0.73 0.75 0.61 0.93 0.88 0.90 Myroodah Air Crossing 0.52 0.86 0.78 0.85 0.66 0.89 0.91 0.92 temperature Camballin Pool 0.50 0.92 0.59 0.79 0.66 0.90 0.75 0.91 Estuarine pools 0.42 0.86 0.61 0.78 0.63 0.90 0.80 0.80 Freshwater pools 0.51 0.92 0.67 0.82 0.66 0.90 0.83 0.85 River 0.50 0.87 0.69 0.80 0.78 0.90 0.85 0.87 Milli Milli -0.25 0.96 - -0.26 - 1.00 - - Snag Pool -0.47 0.96 - -0.23 -0.66 1.00 - - Langi Crossing -0.53 0.97 -0.25 -0.21 -0.58 0.98 - - Myroodah River Crossing -0.64 0.97 -0.24 -0.32 -0.64 1.00 - - discharge Camballin Pool -0.64 0.98 - -0.43 -0.65 1.00 - - Estuarine pools -0.44 0.96 - -0.25 - 1.00 - - Freshwater pools -0.64 0.98 -0.19 -0.38 -0.64 1.00 - -0.55 River -0.77 0.97 - -0.33 -0.86 1.00 - -0.53 Milli Milli - - - -0.15 - - - - Tidal height Snag Pool - - - -0.26 - - - - Langi Crossing - - 0.16 -0.24 - - - -
Water temperatures in King Sound and its tributaries (excluding the Fitzroy
River) ranged from between 19.2 (June/July) and 30.4°C (October) (Fig. 2.7). Mean
sea surface temperature in June/July 2007, was 20.4C ± 0.16 (n = 19) within the
Stokes Bay and Point Torment areas, 26.4C ± 0.4 (n = 6) in September 2009, within
the NE King Sound area and 29.2C ± 0.16 (n = 31) in October 2007, within the SE
and SW King Sound and Point Torment areas. No significant difference was found
between surface temperatures within King Sound and the mouths or upper reaches of
53 its tributaries when all sampling periods were pooled (p = 0.974, d.f. = 2). A significant difference was observed between June/July and September/October surface water temperatures (p < 0.001; n = 18 in June/July, n = 37 in
September/October), as well as bottom water temperatures (p < 0.001, n = 9 in
June/July, n = 8 in September/October). No difference was found between surface and bottom water temperatures in June/July (dT = -0.02C m-1 ± 0.02, n = 5) or in
September/October (dT = 0.03C m-1 ± 0.04, n = 5).
2.3.3 Salinity
Surface water salinity increased significantly between June/July and
September/October/November in all monitored pools (p < 0.001, n = 10 to 22), ranging between 0.2 and 38.4 ppt and 0.2 and 0.6 ppt in the estuarine and freshwater pools, respectively (Fig. 2.10a, b). Differences in the salinity of surface and bottom waters in the estuarine and freshwater pools ranged between 0 and 12 ppt and 0 to 0.1 ppt, respectively. Surface and bottom water salinity was significantly different in the lower estuarine pools (i.e. Lower Pelican Pool, Milli Milli and Snag Pool) throughout the dry season (p < 0.001, n = 10 to 22). However, salinity of surface and bottom waters in Langi Crossing was only significantly different during the late dry season
(June/July: p = 1, n = 18; September/October/November: p < 0.001, n = 17). The salinity of bottom and surface waters was not significantly different in the freshwater pools in the early (p = 1, n = 16 to 17) or late dry season (p = 1, n = 11 to 17).
Salinity measurements taken during the dry season in King Sound ranged between 28.5 and 38.0 ppt. Only in the late dry season in Kimbolton Creek (as referred to by the author), which is located in the Stokes Bay area (16º47’34” S,
123º43’25” E), were salinity levels observed to exceed this range (43.7 ppt at 14 km
54 upstream of the creek mouth). In general, salinity in tributaries was found to increase with distance upstream. In three of the larger creeks, at distances of 5 (NW King
Sound area), 8.8 (Fraser River), 8 and 14 km (Kimbolton Creek) upstream of their respective mouths, surface salinity was 3.6, 3.3, 2.3 and 7.6 ppt, respectively, higher than at the creek mouths. Surface water salinity levels significantly increased between June/July and September/October (p < 0.001; n = 18 in June/July, n = 30 in
September/October), when pooling all King Sound sites. No substantial difference was found between surface and bottom water salinities in either June/July (dS = -0.01 ppt m-1 ± 0.02, n = 5) or September/October (dS = -0.03 ppt m-1 ± 0.02, n = 5).
55 40 Surface water Bottom water
30
20
Salinity (ppt) Salinity
10
0
40
30
20
Salinity (ppt) Salinity
10
NS 0
Milli Milli Snag Pool King Sound Lower Pelican Telegraph PoolLangi Crossing Camballin Pool Myroodah Crossing
Sample site
Fig. 2.10. Mean (+ s.e.) salinity of monitored sites in nearshore waters of the Fitzroy River and King
Sound in (a) June/July and (b) September to November.
56 2.3.4 Turbidity and light attenuation
Turbidity within the estuarine pools of the river ranged between 12.1 ntu during times of no tidal action, and 757 ntu at the peak of high tide in the dry season (n = 19). In contrast, turbidity in the freshwater pools ranged between 2.61 to 7.45 ntu (n = 15) in the dry season. Turbidity in the mud and silt rich nearshore waters of the SW King
Sound area ranged between 208 and 269 ntu when sampled in July 2008 (n = 14).
Turbidity readings in King Sound environments with a more sand-rich substrate, such as the NW King Sound area and the Point Torment area, were observed to be ~20 to
63 ntu (n = 9) in July 2008 and September 2009.
Short-term (~36 h) monitoring periods demonstrated light attenuation to be substantially greater in estuarine pools than in freshwater pools (Fig. 2.11). Pendant loggers deployed in June, recorded a maximum of 23 % (hours 12 and 13 were not documented due to system malfunctions) and 13 % of available light intensity to have penetrated depths of 0.5 m in Myroodah Crossing and Snag Pool, respectively. A maximum of 8 % and <1 % of available light intensity was detected at a depth of 1.5 m in Myroodah Crossing and Snag Pool, respectively.
Long-term (months) light intensity data was limited as only three of the eight loggers, which were deployed in intervals of 0.5 m on moorings in Myroodah
Crossing and Telegraph Pool in November 2008, were recovered (one at a late dry season depth of 0.5 m in Telegraph Pool, and two in Myroodah Crossing at late dry season depths of 1.5 and 2.0). Wet season data from Myroodah Crossing demonstrated a strong negative correlation to exist between light intensity and stage height (ρ = -0.81, n = 144; Fig. 2.12).
57 10000
Telegraph Pool (0.5 m) Myroodah Crossing (1.5 m) 8000
6000
4000
Light intensity (lux) intensity Light
2000
0
25-Nov-08 29-Nov-08 03-Dec-08 07-Dec-08 11-Dec-08 15-Dec-08 Date
Fig. 2.11. Light intensity (lux) in Telegraph Pool (0.5 m) and Myroodah Crossing (1.5 m) at the start of the 2009 wet season.
3500 18 Light intensity (lux) Stage height (m) 3000
16 2500
2000 14 1500
Stage height (m) height Stage
Light intensity (lux) intensity Light 1000 12
500
0 10 01-Dec-08 01-Jan-09 01-Feb-09 01-Mar-09 01-Apr-09 Date
Fig. 2.12. Logged light intensity (lux) in Myroodah Crossing (at a depth of 1.5 m) during the 2009 wet season, in relation to the stage height recorded at the Camballin Barrage (~38 km upstream of
Myroodah Crossing).
58 2.3.5 Dissolved oxygen and pH
Dissolved oxygen in surface waters was found to decrease between June/July and
October/November in the estuarine and freshwater pools. Dissolved oxygen decreased from a mean of 10.0 ± 0.2 (n = 5) to 7.5 ppm ± 0.3 (n = 7) in the estuarine pools. Dissolved oxygen decreased from a mean of 9.2 ± 0.4 (n = 4) to 7.3 ppm ± 0.6
(n = 2) in the freshwater pools. Differences between surface and bottom water DO was small in the estuarine (0.1 ppm ± 0.2, n = 3; depth <2.7 m) and freshwater pools
(0 and 0.05 ppm m-1) in June/July, and in the estuarine pools in November (0.2 ± 0.3, n = 7; depth <2.5). Mean DO in the SW King Sound area in the early dry season was
9.35 ppm ± 0.22 (n = 8) in surface waters. The mean difference in DO between surface and bottom water in this area was 0.02 ppm m-1 ± 0.02 (n = 5).
The few pH measurements recorded during this study showed a decrease in pH between June and November 2008 in the freshwater (8.24 and 8.1 to 8.09 and
7.64) and estuarine pools (8.3 ± 0.1 [n = 4] to 7.1 ± 0.1 [n = 7]). However, data provided by the Department of Water, Government of Western Australia demonstrated a significant increase in pH at their Camballin Barrage and Looma stations (i.e. freshwater areas) between May/June/July (8.0 ± 0.05, n = 18) and
October/November (8.5 ± 0.03, n = 12) (p < 0.001, d.f. = 11). Wet season measurements of pH were scarce and consisted of two records at the Camballin
Barrage station of 6.1 (December 1990) and 7.4 (February 1998), and one at the
Looma station of 7.5 (February 1998). Mean recorded pH was 8.3 ± 0.03 (n = 7) in the SE and SW King Sound areas in July 2008.
59 2.3.6 Tidal cycle
High tide ranged between 6.64 and 11.74 m during the study period. The mean lag time of high tide between the Derby Jetty and Snag Pool was 3.1 h ± 0.1 (n = 11), although tides were first observed at Snag Pool 1.7 h ± 0.1 (n = 9) after high tide at the jetty. Tides of ~9.5, ~10.5 and ~11 m were observed at Snag Pool, Telegraph
Pool and Langi Crossing, respectively. Slack tide at Snag Pool was measured to occur for ≤0.25 h.
2.4 Discussion
Knowledge of an environment including the dynamics of its components and at what spatial and temporal scales these change, are central in understanding how species are influenced by, and use habitats. Environmental monitoring of the Fitzroy River and
King Sound revealed a high degree of spatial and temporal habitat heterogeneity in salinity, water temperature, turbidity, light attenuation and DO, which are referred to in this study as secondary variables. Many of the temporal changes in secondary variables appeared to be associated with, and likely influenced by more than one primary variable, including air temperature, tidal action as well as river discharge and stage height. However the strength of the relationship between these variables was shown to change with space and time. Differences in hydrologic geometry and substrate composition also appeared to vary the strength of the relationship between primary and secondary variables.
60 2.4.1 Temperature
Water temperature was found to be one of the most inconstant environmental components explored, and greatly changed between macrohabitats (i.e. regions and pools). Smaller bodies of water were found to be less thermally stable than larger ones. This was evident in comparing seasonal and diel temperature ranges of the shallow estuarine pools with the deeper freshwater pools, and of the river with King
Sound. Variances in the temperature range between these locations can be in part explained by differences in their hydrological geometry, as larger surface areas allow for greater rates of energy exchange and shallower depths require less energy to heat
(LeBlanc et al. 1997; Malcolm et al. 2004). Riparian shading is also known to reduce thermal variation by screening out incoming solar radiation and reducing the export of long-wave radiation (Larson and Larson 1996; LeBlanc et al. 1997). It is reasonable to suspect that the differences in diel temperature ranges between the freshwater pools and the more barren estuarine pools (Storey et al. 2001), may also be due to differences in vegetated cover. However, similarities in the temperature profiles of
Langi Crossing, an estuarine pool, and the freshwater pools, which are similar in depth, suggests that vegetated cover may be less influential on changes in water temperature than the hydrological geometry of a pool.
Differences between the temperatures of microhabitats, namely bottom and surface waters, were also found to change between the different regions, in part due to differences in the frequency and strength of mixing forces. The frequent occurrence of large tides in King Sound has been shown to prevent stratification of the water column (Wolanski and Spagnol 2003); similar to what was observed in this study. A decrease in the frequency and strength of tides appeared to result in an increase in the thermal stratification of the water column. This was evident in the significant
61 difference between surface and bottom water temperatures in Langi Crossing, where tides rarely penetrate, as opposed to the more mixed lower estuarine pools, where tides are a more frequent occurrence. Development of a stratified water column in the river with progression of the dry season can also be reasoned to be caused by the reduction of mixing from reduced river flow.
Water temperature was also temporally sensitive and followed a fairly predictive seasonal pattern. However, unlike the commonly accepted wet/dry seasonal division used for tropical environments, water temperature within the Fitzroy
River followed a quarterly pattern: early wet season (December to March), late wet season (April to May), early dry season (June to August) and late dry season
(September to November). It was during the late wet and late dry season that temperature varied the most. Seasonal and daily differences in water temperature have been attributed to numerous physical variables including river discharge (Webb and Zhang 1997; Sinokrot and Gulliver 2000), ground water intrusion (Webb and
Zhang 1997; Arscott et al. 2001), air temperature (Pilgrim et al. 1998; Langan et al.
2001; Malcolm et al. 2004; Webb and Nobilis 2007) and radiant energy (Evans et al.
1994; Webb and Zhang 1997; Sinokrot and Gulliver 2000; Malcolm et al. 2004). By analysing water temperature by season rather than by year, this study was able to more closely demonstrate the complex relationship between water temperature and the primary variables listed above.
Air temperature had the strongest relationship with water temperature. This was likely to be a result of both variables being driven by a common denominator, namely radiant energy (Evans et al. 1994; Larson and Larson 1996; Webb and Zhang
1997; Sinokrot and Gulliver 2000), and due to the effect of air temperature on water temperature (Pilgrim et al. 1998; Langan et al. 2001; Malcolm et al. 2004; Webb and
62 Nobilis 2007). However, deviations in the strength of this correlation were observed to occur at multiple temporal scales. Periodic discrepancies between air and water temperature have been shown to occur in times of increased river flow and snowmelt
(Smith and Lavis 1975; Webb and Zhang 1997; Sinokrot and Gulliver 2000). In this study the correlation strength between air and water was at its lowest during the early wet season when discharge is typically greatest and the relationship between discharge and water temperature is growing. The significant increase in flow during the wet season was likely to have caused water temperature to deviate from air temperature, as the greater volume of water would have altered the river’s thermal properties (Sinokrot and Gulliver 2000).
Seasonal analyses also showed water temperature to develop a relationship with tidal action during the late dry season. Although it was surprising to observe that this relationship was not stronger, having observed the decrease in water temperature maxima in the estuarine pools during the presence of colder tidal waters. A consistent increase in water temperature during the late dry may have diminished the apparent relationship between these variables in the statistical analysis. Filtering out this seasonal effect in further analyses may show tides to have a stronger relationship with water temperature.
2.4.2 Salinity
Salinity significantly increased during the dry season in the estuarine pools, due primarily to the concurrent decrease of freshwater input and continuous input of marine tidal water. Wolanski and Spagnol (2003) also observed a similar pattern in the upper King Sound during the dry season. Wolanski and Spanol (2003) suggested that the increase in salinity is caused by the inability of freshwater input to make-up
63 the water level balance being shifted by increasing evaporation rates, leaving only the saline ocean water to fill the difference.
Salinity was also observed to vary between bottom and surface waters in the lower estuarine pools during the dry season. The formation of a salt gradient between surface and bottom waters is often due to decreases in mixing forces, like river flow, wind and/or tides (Simpson et al. 1990; Peters 1997). The periodic lull in tidal action in the estuarine pools allowed haline stratification. This was present as a small salt wedge in the early dry season when river flow maintained fresh surface waters.
However, Peters (1997) found that even a well-developed salt gradient in a river estuary can be reduced or eliminated temporarily by tidal actions (Simpson et al.
1990; Peters 1997). It is likely that stratification of the estuarine pools is also periodically reduced or eliminated as a result of the frequent penetration of the macrotides. Conversely, the results from King Sound, which is frequently influenced by macrotides, showed constant mixing to produce a homogenous water column throughout the year. Whereas minimal salinity levels (<1 ppt) in the freshwater pools prevented the detection of a significant difference in surface and bottom waters.
2.4.3 Turbidity and light attenuation
Turbidity in King Sound was substantially greater in the shallow upper reaches (from the Fraser River in the SW King Sound area to Kimbolton Creek in the Stokes Bay area) than deeper offshore waters and locations closer to the mouth of the sound (e.g. the NE King Sound area, offshore site at the tip of the Point Torment area). Wolanski and Spagnol (2003) also observed a similar trend and reported on the presence of a turbidity maximum zone in the upper King Sound. Wolanksi and Spagnol (2003) concluded that while tidal pumping and biological filtering of offshore muddy marine
64 snow had generated this anomaly, it was the wind-driven waves that maintained it through at least the dry season. Substrate in areas of relatively low turbidity appeared to have greater sand to mud/clay ratio than those in the turbidity maximum zone.
Once mixed, the heavier sediments in these regions would settle out faster (Foster et al. 1992). Turbidity levels cannot be commented on for the wet season as samples were not taken during this period. However, increased discharge during the wet season is likely to increase the range of turbidity plumes from tributaries.
Turbidity in the Fitzroy River was found to decrease with distance traveled upstream during the dry season, presumably due to the reduction of tidal influence and increase in sediment mass. Cyrus and Blaber (1992) observed a similar pattern between May and November in the Embley Estuary, but noted a reverse turbidity gradient throughout the wet season. A substantial increase in turbidity in the wet season was also evident in the Fitzroy River through satellite and land-based photographs. This was further substantiated by the significant drop in light intensity observed in the wet season, which was likely due to light being scattered and absorbed by an increase in suspended sediments (Jerlov 1968; Toro 1978; Kirk 1985;
Bhargava and Mariam 1991). Using light intensity as an indirect measure of turbidity, wet season turbidity levels in the freshwater pools of the Fitzroy River were found to be even greater than that observed in the estuarine pools in the dry season.
The increase in suspended sediment during the wet season is assumed to be a result of increased river flow (i.e. suspension of sediments), as well as sediment import from run-off and lateral transport from the floodplain (Junk et al. 1989; Taylor 1999).
Erosion is relatively minimal in the freshwater pools of the Fitzroy River (Taylor
1999) and not as likely to have contributed to the increased suspended sediment levels. Suspended sediment in the freshwater pools appear to dissipate within a
65 couple weeks after the final wet season ‘flood event’ (for the purposes of this study, a flood event is defined as: a sudden substantial increase in river discharge, resulting in a >0.5-m rise in stage height).
2.4.4 Dissolved oxygen and pH
Dissolved oxygen is a critical component in riverine and estuarine systems and is commonly reduced by increases in algal photosynthesis (Chaudhury et al. 1998), sediment oxygen demands (Boynton and Kemp 1985; Cowan et al. 1996; Chaudhury et al. 1998), respiration (Kemp and Boynton 1980; Stanley and Nixon 1992; Kemp et al. 1992; Chaudhury et al. 1998) and nitrification (Balls et al. 1996; Chaudhury et al.
1998; Dai et al. 2006). In rivers like the Fitzroy, decreases in DO from reducing factors are exasperated during the late dry season due to a reduction in oxygen- absorption of warmer water and decreased mixing at the air-water interface (from reduced flows) (Stanley and Nixon 1992). Such circumstances have led to anoxic bottom waters in numerous eutrophic estuaries (Kemp et al. 1992; Balls et al. 1996;
Li et al. 2002; Abril et al. 2003; Dai et al. 2006). Although DO levels did decrease during the dry season, likely due to decreased river flow, elevated water temperature and increased algal growth, levels remained well above hypoxic.
Occurrence of similar DO levels in surface waters of the different regions was unexpected as oxygen is more readily dissolved in water with lower concentrations of dissolved or suspended solids. Dissolved oxygen was expected to be lower in the more saline and turbid King Sound and estuarine pools. The extreme tidal action in the relatively shallow environments of King Sound and estuarine pools was likely to have allowed for a greater mixing of oxygen into surface waters (Stanley and Nixon
1992). This tidal mixing is thought to have also been responsible for the lack of a
66 difference between DO in surface and bottom waters. Haline stratification, which is seen in salt wedge estuaries, often greatly reduces the exchange of DO between surface and bottom waters (Sanford et al. 1990; Kemp et al. 1992). However, tide induced advection and homogenisation of the water column in the estuarine pools and
King Sound potentially facilitated exchange of DO between these two microhabitats
(Sanford et al. 1990; Kemp et al. 1992).
Unlike DO, pH in the freshwater pools (documented by the Department of
Water, Government of Western Australia) exhibited a general increase between April and November, and a decrease between December and March. This rise in pH likely stems from an increase in salinity, specifically calcium salts, which are left behind as pool water evaporates (Wong 1979; Prathumratana et al. 2008). Alternatively, this change in pH may be showing evidence of the chemical make-up of the groundwater from the alluvial aquifers, which are the primary water sources of the pools during the late dry season (Lindsay and Commander 2005).
The pH trend observed by this study in the estuarine pools (decreasing from
8.5 in June/July to 7.4 in October/November) is not well understood and requires further analysis. The pH in the estuarine pools was expected to increase through the dry season due to the intrusion of marine water, which is characteristically alkaline, and the significant increase of salinity in the estuarine pools (Wong 1979; Fuentes-
Yaco et al. 2001). The reason for the observed decrease in pH is unknown and further sampling is needed to confirm and explain these results.
2.4.5 Conclusion
This study demonstrated the Fitzroy River and King Sound environments to be highly variable through space and time. Analyses of air and water temperature, river
67 discharge and stage height, tides and light intensity (to a certain degree) in the river were possible through the use of temperature and light data loggers as well as from data supplied by the Department of Water, Government of Western Australia and the
Commonwealth of Australia Bureau of Meteorology. However, only general seasonal patterns between the early dry and late dry seasons were commented on for turbidity,
DO, pH and salinity, as well as water temperature in King Sound, because of the limitations from opportunistic sampling. However, by understanding the patterns of the primary variables and their effects on secondary variables in similar systems, inferences could be made about the temporal patterns of turbidity, DO, pH and salinity. Using these data, this study was able to document the temporal patterns and spatial differences of these variables and the scales in which significant change occurs.
As hypothesised, the secondary variables such as water temperature were closely related to, and appeared to be influenced by the temporal patterns of multiple primary variables, as was observed with water temperature and air temperature, tidal action and river discharge. In several cases, the strength of the relationship between primary and secondary variables changed over time. For example, water temperature typically followed a pattern similar to that of air temperature, but deviated from this pattern during specific seasons when the correlation strength between other variables, such as tidal action and river discharge, and water temperature increased. The degree to which these temporal changes in environmental variables occurred appeared to be influenced by the spatial heterogeneity of the primary variables (e.g. hydrologic geometry, tidal action, substrate composition). Spatial differences in primary variables created variances in water temperature, salinity and turbidity/light attenuation between the regions. This was evident in the difference in daily
68 temperature range of the estuarine and freshwater pools of the river, with daily variations being much greater in the wider, shallower and less shaded estuarine pools.
A knowledge of how primary and secondary variables differ spatially and temporally and at what scales are likely to be important for assessment of the impacts of anthropogenic or climatic change on this environment. In regards to the subject of this thesis, this information can be used to determine how changes in primary variables may indirectly affect the local aquatic fauna. As an example, the transformation of the Fitzroy River from a seasonally to annually flowing river by flow regulating dams, would likely reduce the salinity increase in the estuarine pools during the dry season. This would potentially make the estuarine pools more suitable year round for juvenile P. microdon (see Chapter 4). However, this would also prevent access to this area by marine animals, like the dwarf sawfish (Pristis clavata) found to use this area in the late dry season (Whitty et al. unpublished data).
69 Chapter 3: Spatial and temporal changes in relative
abundance, distribution, sex ratio, natural mortality and growth of the freshwater sawfish Pristis microdon
and northern river shark Glyphis garricki in the
Fitzroy River and King Sound, Western Australia
3.1 Introduction
Simple catch and catch per unit effort (CPUE; catch data used in association with fishing effort) data can be useful in documenting the demography and biology of a population (Stevens and McLoughlin 1991; Cortés 1998; Simpfendorfer 2000;
Heithaus 2001; Merson and Pratt 2001; Pikitch et al. 2005) and development of effective management strategies for fishes (e.g. Simpfendorfer 2000) and can be used to document how a population uses a habitat. For example, presence and relative abundance data of sex and size classes in a population can be useful in determining the general purpose (e.g. pupping ground, nursery, adult habitat) of a habitat and timing of its use. Also, as life history parameters, the “phenotypic expressions of the interaction between the genotype and the environment” (Begg 2005; White and
Sommerville 2010), are regulated by environmental pressures, they may be useful indicators that reflect the influences of environmental variables on a population.
Unfortunately, establishing a statistically robust data set that accurately represents a population of critically endangered elasmobranchs, such as the freshwater sawfish
(Pristis microdon) and northern river shark (Glyphis garricki), can be difficult due to their rarity. In such circumstances, long-term sampling and/or collaboration between
70 studies can aid in increasing sample sizes (Simpfendorfer 2000; Francis and Duffy
2002; Theberge and Dearden 2006).
Pristis microdon and G. garricki populations in northern Australia are believed to be in decline and/or limited in number (Pogonoski et al. 2002; Pogonoski and Pollard 2003; Stevens et al. 2005). As a result, these species were listed as
Critically Endangered on the IUCN Red List of Threatened Species (Pogonoski and
Pollard 2003; Compagno et al. 2006) and as Endangered (G. garricki) and Vulnerable
(P. microdon) on Australia’s Environment Protection and Biodiversity Conservation
(EPBC) Act (Department of Sustainability, Environment, Water, Population and
Communities 2011a, b). Until recently, very little has been known about these species, which has made assessment and management of Australian populations difficult (Stevens et al. 2005). A handful of recent (post 1990) field studies have explored the demography, life history and habitat use of P. microdon (Tanaka 1991;
Taniuchi et al. 1991; Thorburn et al. 2003, 2007; Peverell 2005, 2008; Thorburn
2006; Field et al. 2008) and G. garricki (Tanaka 1991; Taniuchi et al. 1991; Thorburn et al. 2003; Thorburn and Morgan 2004; Thorburn 2006; Field et al. 2008) in northern
Australia. Unfortunately, limited sample sizes (n = 5 to 79 P. microdon and 1 to 10
G. garricki) and study periods (months to a few years) have limited the robustness of the results of these investigations. This has left uncertainty in the understanding of growth rates and distributions and no published analyses of relative abundance or mortality. As such, the aims of this study were to document annual and seasonal
CPUE of P. microdon and G. garricki, growth and mortality rates of P. microdon as well as demography of P. microdon and G. garricki within Western Australia. A five-year sampling regime targeting P. microdon and G. garricki was conducted in the
Fitzroy River and King Sound, Western Australia, to accomplish these aims. Sample
71 sizes were increased when needed, by compiling recorded data with previous research
(commenced in 2002) of these populations (Thorburn and Morgan 2004; Thorburn
2006; Thorburn et al. 2007) to provide the most accurate results. This compilation represents the longest running and largest dataset of these species to date.
3.2 Material & Methods
3.2.1 Study Site
Sampling of P. microdon and G. garricki was conducted in the upper (i.e. southern)
King Sound and in the lower 160 km of the Fitzroy River in northern Western
Australia (Fig. 3.1). The aquatic environments of the study site undergo extreme daily, seasonal and annual fluctuations during transitions between wet (December to
April) and dry (May to November) seasons and during large (>11 m) mixed-tides.
The Fitzroy River is a tropical river that flows through the western Kimberley and discharges into King Sound (Fig. 3.1). River discharge and stage height is highly seasonal. Stage height in the wet season occasionally peaks at >10 m above the dry season levels. In the dry season, the river exists as an extensive series of isolated pools intermittently connected by shallow (<0.5 m in depth) stretches in the river (i.e. runs). Large mixed-tides (9 to 12 m) periodically inundate the lower 17.5 km of the river, connecting pools within this region and creating a well-mixed and turbid water column (Chapter 2). Upstream pools are fresh (<1 ppt) and become thermally stratified with progression of the dry season (Chapter 2).
72
Fig. 3.1. Sampled sites (red dots) and study areas (shaded) in the Fitzroy River and King Sound
between 2005 and 2009. Green shading = NW King Sound area, orange shading = SW King Sound
area, yellow shading = SE King Sound area, red shading = Point Torment area, blue shading =
Stokes Bay area.
Nearshore environments of the upper King Sound and its tributaries are typically characterized by extreme low visibility. These nearshore waters average
<10 m in depth, while King Sound as a whole has mean and max depths of 18 and 50 m, respectively (Wolanski and Spagnol 2003). During low tides, most of the more sand-rich (with mud and silt inter-dispersed) nearshore areas and the muddy tributaries are often exposed. Larger creeks (10 to 20 m in width) are separated from
73 other waterways by dense mangrove forests and narrow with progression inland.
Larger tributaries, like the Robinson and May River estuaries are wide (100s of metres) low-relief channels that are lined by low-lying mud flats.
3.2.2 Sampling
Sampling for P. microdon and G. garricki occurred in the Fitzroy River between June
2005 and November 2009, and in King Sound between June 2007 and September
2009 (Fig. 3.1). Sampling consisted of over 3329 h of 20-m net sets at 11 different sites in the lower 160 km of the Fitzroy River, and over 1342 h of 20-m net sets at 25 sites in King Sound. Although 50 and 70-m nets were deployed, 1 h sets of these nets were transformed to 2.5 or 3.5, respectively, 20-m net sets in CPUE calculations to simplify matters. Sites sampled in King Sound were grouped into five areas: North-
West (NW), South-West (SW), and South-East (SE) King Sound, as well as Point
Torment and Stokes Bay) (Fig. 3.1). Nets that were deployed in the river consisted of
20 m of 51, 76, 102, 152 and 203-mm stretched monofilament gillnets. Nets deployed in the sound consisted of 20 m of 152 and 203-mm, 70 m of 102-mm, as well as 50 m of 102 and 178-mm stretched monofilament gillnets. Size of net mesh deployed varied between all years of sampling, including 2002 to 2004 sampling conducted by
Thorburn et al. (2007) (Table 3.1). Only 102 and 152-mm meshes were deployed in all years, and together were used in 50 to 100% of net sets per year, with exception of
2003 (Table 3.1). In 2003, 51-mm mesh was used for 73.4 % of net sets (Table 3.1).
Nets were placed in depths of one to six metres in the river and one to eight metres in
King Sound. Nets were checked at one to two hour intervals unless activity was observed in the net, at which time they were checked immediately. Checking nets prematurely may have introduced bias into relative abundance estimates, but was
74 determined essential in order to minimise the impact of capture on these critically endangered species. Soak time was extended between four to six hours between
00:00 and 06:00 h depending on bycatch levels. Sampling was also conducted by hand-lines used in conjunction with a variety of bait including lesser salmon catfish
(Neoarius graeffei) and mullet (Mugilidae spp.).
Table 3.1. Percentage of yearly 20-m net sets for each respective gillnet mesh size deployed
Year 51 mm 76 mm 102 mm 152 mm 203 mm 2002 38 62 2003 73 9 18 2004 23 14 49 14 2005 23 23 27 27 2006 8 17 50 26 2007 18 57 26 2008 1 21 78 2009 40 52 8
Pristis microdon captured in the Fitzroy River, were typically moved to the shoreline to be measured. The majority of the sawfish, including the spiracles and/or mouth, remained submerged during handling. Occasionally P. microdon <1200 mm total length (TL), were measured and tagged on board a research vessel. This was done to decrease the time between capture and release. In King Sound, relatively small (<2000 mm TL) elasmobranchs were measured and tagged on a research vessel.
A 25-mm diameter hose was inserted into the mouth of onboard specimens to allow ambient water to be pumped through the gills of a fish when more than TL was to be measured. Elasmobranchs too large to land (>2000 mm TL) were measured and tagged over the side of the vessel to prevent damage to the fish and potential injuries to researchers. Upon capture, elasmobranchs were inverted to induce tonic immobility (Henningsen 1994; Watsky and Gruber 1990).
75 Individually numbered Rototags (Dalton Supplies, Woolgoolga, New South
Wales, Australia) were fitted to the first dorsal fins of P. microdon and G. garricki, following procedures similar to that described by Heupel et al. (1998). Deployed
Rototags were labeled with contact information to allow fishers to report recaptures of these individuals. In addition, road-side posters had previously been installed at key sites (e.g. boat ramps) instructing fishers on what to do if a tagged P. microdon was recaptured.
3.2.3 Data analysis
Statistical analyses were conducted using SPSS V15 (SPSS, Chicago, Illinois, USA) or SigmaPlot 11.0 (SPSS, Chicago, Illinois, USA). Data from Thorburn and Morgan
(2004), Thorburn (2006) and Thorburn et al. (2007) were included (with permission from the authors) in sex ratio, CPUE and growth analyses to increase sample sizes and to establish inter-annual patterns over a greater period. Sampling during this previous research was conducted during months and at sites (within the lower 160 km of the Fitzroy River) that were to this study. Maps were produced using ArcGIS 9.3.1
(ESRI, Redlands, California, USA).
Sex ratios were calculated for P. microdon captured in the Fitzroy River for the years 2002 through 2009, and for all years combined. Annual sex ratios were calculated separately for young of the year (YOY), <2400 mm TL P. microdon (i.e. the size range of male P. microdon in the river) and <2800 mm TL P. microdon (i.e. the size range of female P. microdon in the river) to explore if sex ratios varied with size. Classification of individuals as YOY was based upon the presence of an umbilical scar and/or TL. Classification of P. microdon to the YOY size class was based on findings by Thorburn et al. (2007) and Peverell (2008), and generally
76 consisted of individuals <1100 mm TL. A chi-square test was run on each group to determine if the sex ratios significantly varied from a 1: 1 ratio.
Catch per unit effort was calculated for P. microdon and G. garricki using 100 h of 20-m net as the standard unit of effort. Catch per unit effort in the river was pooled for all net mesh sizes as the majority of sets consisted of the similar 102 and
152-mm meshes, and as selectivity of mesh sizes appeared broad with P. microdon captures (i.e. all sizes classes were captured in all deployed mesh sizes). Net selectivity was not statistically analysed due to the small number of P. microdon captured in different mesh sizes that were deployed in similar conditions (e.g. time of day, season, pool, depth). Nets deployed under different conditions were not pooled for analyses as this would likely have introduced bias into the results. Related results should be treated with caution until net selectivity can be demonstrated.
In CPUE analyses, primary months of sampling: June, July, October and
November were pooled into two separate periods representing the early dry season
(June and July) and the late dry season (October and November) (see Chapter 2).
Catch per unit effort resulting from minimal sampling efforts in other months was excluded from analyses to ensure independence between seasons (exclusion of data resulted in minor changes in only early 2004 and late 2008). However, sampling efforts in late May and September 2009 were included in analyses, as sampling was not possible during the other months of this year. Total CPUE per season was analysed in place of mean CPUE per season, as the large number of sets with zero captures would have biased results. Seasonal (i.e. early and late dry seasons) changes in total CPUE of YOY and >0+ P. microdon were tested for significance using a paired t-test. In addition, the percentage of positive captures (i.e. percentage of seasonal effort that resulted in capture of at least one P. microdon) was graphically
77 illustrated and visually compared to that of total CPUE. If similar patterns between these two figures were observed, it would suggest that the calculated total seasonal
CPUE were not heavily biased by outliers.
Discharge and stage height are some of the most extreme abiotic variables within the Fitzroy River (Chapter 2). It was hypothesised that the extreme changes in discharge and/or stage height would influence the recruitment and/or survivorship of
P. microdon within the Fitzroy River, as has been reported for other estuarine and freshwater fishes (Mills and Mann 1985; Drinkwater and Frank 1994). A Pearson product moment correlation was run between CPUE in the early and late dry season and total wet season stage height (total stage height was the sum of daily maximum stage heights for the analysed period) to test this hypothesis. Discharge and stage height data were measured at Willare Station (see Fig. 2.2) in January to April. River stage height and discharge were collected and made available by the Department of
Water, Government of Western Australia. Additionally, a linear regression was run for CPUE and total wet season stage height to investigate the influence of stage height on the percentage of YOY to move upstream into the studied freshwater pools (~120 to 160 km upstream of the river mouth) in the early dry season. Late dry season numbers were not included in this analysis as upstream movement is unlikely during this time, as certain runs in the river are typically dry by this period.
Natural mortality (M) of juvenile P. microdon was estimated using multiple indirect methods to provide insight into changes between the early and late dry season
CPUE. These included three age-independent (Pauly 1980; Jensen 1996) and three age-dependent methods (Peterson and Wroblewski 1984; Chen and Watanabe 1989;
Lorenzen 1996) (Table 3.2).
78 In the Peterson and Wroblewski and Lorenzen methods, wet weight (Wt) was used in place of dry weight. Wet weight was found to produce a better estimate of M for numerous shark species (Cortés 2002). Weight-at-age used for these methods was calculated using an observed wet Wt (g) to TL (mm) relationship (Wt = 3E-07TL3.2831) derived from data from this study and Thorburn et al. (2007) and length-at-age estimates used for juvenile (0+ to 5+ year old) P. microdon. Length-at-age estimates were estimated from vertebral analysis (Thorburn et al. 2007; Peverell 2008) and visual analysis of length-frequency data from this study.
Parameters for the von Bertalanffy growth function (VBGF) used in the Pauly
(1980) and Jensen (1996) methods were obtained from output values of the LFDA
V.5 package (discussed below), as well as from back calculated data reported by
Peverell (2008). As t0 (estimated age at length 0) produced from the LFDA package can be unrealistic and is constrained between 0 and -1, only VBGF parameters from
Peverell (2008) were used for the Chen and Watanabe method, which uses t0 as a variable. Water temperature used in the Pauly (1980) method, which assumes the concept that water temperature will influence mortality rates, was the mean annual surface temperature of Myroodah Crossing, Fitzroy River (28 °C) (see Fig. 3.1,
Chapter 2). River temperature was determined to be more appropriate than that of
King Sound, as this study focused on juvenile mortality. It was not possible to estimate fishing mortality (F), but F is assumed to be comparatively low for juvenile
P. microdon as commercial fishing is not allowed within the Fitzroy River. Natural mortality estimates were used to calculate annual survival (S = e-M) of P. microdon.
79 Table 3.2. Methods used for calculating natural mortality (M) estimates
L, to, von Bertalanffy growth parameters; T, average annual temperature in °C; Wt, wet body weight
in grams; t, age in years
Cited by Method
Pauly (1980) log M = -0.0066 - 0.279 log L + 0.6543 log k + 0.4634 log T -0.25 Peterson and Wroblewski (1984) MW = 1.92 W -k(t - to) Chen and Watanabe (1989) Mt = k/1 - e Jensen (1996) M = 1.5 k Jensen (1996) M = 1.6 k -0.289 Lorenzen (1996) Mw = 2.53 W
Pristis microdon VBGF parameters were estimated using multiple length-
frequency analysis methods incorporated into the LFDA V.5 software package
(Kirkwood et. al. 2001, www.fmsp.org.uk/Software.htm). Analyses included a
projection matrix (PROJMAT) method (Rosenberg et al. 1986; Basson et al. 1988)
and ELEFAN 1 method (Pauly 1987). In addition, the Hoenig and Choudary
Hanumara seasonal growth curve (Hoenig and Choudary Hanumara 1982) was used
in conjunction with both methods to analyse the potential of seasonal growth. A
generated score helped to determine the best fit of growth curves to the data. Higher
scores represented better fits but were not comparable between methods. The fit of
the produced growth curves to the data was visually assessed to compare methods.
3.3 Results
3.3.1 Pristis microdon
3.3.1.1 Catch composition and distribution
Ninety-four different P. microdon were captured between July 2005 and September
2009 (three in 2005, one of which was a recapture of a 2004 tagged fish, 15 in 2006,
80 one of which was a recapture of a 2003 tagged fish, 37 in 2007, 18 in 2008 and 21 in
2009), in the Fitzroy River (Fig. 3.2, 3.3, Appendix I). An additional P. microdon was found dead on the shores of Telegraph Pool in November 2008. Females ranged between 789 and 2750 mm TL. Males ranged between 763 and 2210 mm TL (Fig.
3.3, Appendix I).
In King Sound, one P. microdon was captured in October 2007 in the Point
Torment area (Fig. 3.2) and measured 2580 mm TL. This was the only captured male
P. microdon with semi-calcified claspers. This individual was captured ~5 km from the shoreline on a large and partially submerged sandbar that bordered a relatively deep (~20 m) channel.
Pristis microdon was captured in the Fitzroy River via gillnet in water that ranged in temperature between 17.3 to 31.4oC, and by hook and line in water temperatures between 21.5 and 29.4oC. Salinity ranged from 0.0 to 20.1 ppt at the time of capture of these animals. Surface water temperature in King Sound was
29.0oC and salinity was 35.6 ppt at the time of capture of the single P. microdon in the
Point Torment area.
81
b 16°0'S Western Australia
Robinson River )
King Sound 17°0'S Meda River )) May River Fraser River ) ) ) Telegraph Pl
) Geike Gorge
18°0'S )) ) Lwr Barrage Pl ) )) ) ) ) ¸ Fitzroy River ) 0 1020 40 60 80
19°0'S Kilometers 123°0'E 124°0'E 125°0'E 126°0'E This study ) Previous study
Fig. 3.2. Pristis microdon capture locations in 2005 to 2009 (red dot), and in 2002 to 2004 (Thorburn
et al. 2007) (blue square).
82 12 10 Females 2002 2006 8 Males 6 4 2
Number of fish of Number 0
12 10 2003 2007 8 6 4 2
Number of fish of Number 0
12 10 2004 2008 8 6 4 2
Number of fish of Number 0
12 10 2005 2009 8 6 4 2
Number of fish of Number 0
600 800 600 800 1000 1200 1400 1600 1800 2000 2200 2400 2600 2800 3000 1000 1200 1400 1600 1800 2000 2200 2400 2600 2800 3000 Total length (mm) Total length (mm)
30 25 2002 - 2009 20 15 10
Number 5 0
600 800 1000 1200 1400 1600 1800 2000 2200 2400 2600 2800 3000 Total length (mm)
Fig. 3.3. Pristis microdon total length-frequency histograms for 2002 to 2009.
Sex ratios (male: female) of juvenile P. microdon in the Fitzroy River was highly variable between 2002 and 2009, ranging between 1: 0.47 and 1: 3.17 in each year, and averaged 1: 1.5 ± 0.30 (mean ± s.e.) (Table 3.3). Males only outnumbered females (1: 0.47) in 2002 (Table 3.3, Fig. 3.3). No P. microdon were tagged in 2002, and individuals may have been counted more than once if recaptured, potentially biasing sex ratio results. It should also be noted that nearly half (7 of 17) of the males
83 recorded in 2002, were found deceased in a single pool in October at the time of sampling. Excluding the seven males and two females found deceased in 2002, the sex ratio for the year was 1: 0.6. Sex ratios of YOY in each year that P. microdon were marked were near equal, with exception of 2008, when females outnumbered males (1: 2.3) (Table 3.3). The sex ratio for YOY was 1: 1.12 and did not significantly vary from the sex ratio 1: 1 (χ2 = 0.23, d.f. = 1) when all years were pooled. Pristis microdon >2400 mm TL, were minimal in most years and only comprised >10% of captured P. microdon in 2003 to 2006. During this period the sex ratios for all juvenile P. microdon was between 1: 1.38 and 1: 3.17. In pooling these years the sex ratio was 1: 2.1 and varied significantly from the sex ratio 1: 1 (χ2 =
7.81, d.f. =1) (Table 3.3). Conversely, sex ratios for P. microdon <2400 mm TL during this period was 1: 0.99 (Table 3.3) and did not vary significantly from a 1: 1 ratio (χ2 = 0.67, d.f. = 1).
Table 3.3. Number of captured male/female Pristis microdon for respective years and size
classes in the Fitzroy River
All, all captured size classes; YOY, young of the year
Year All <2400 mm TL YOY 2002 17/8 15/6 1/2 2003 6/19* 6/8 1/0 2004 8/11* 8/5 1/1 2005 1/2* 1/1 0/0 2006 5/10* 5/5 3/2 2007 18/19 18/19 16/16 2008 9/10 9/10 3/7 2009 10/10 10/9 8/9
* Pristis microdon >2400 mm TL made-up >10 % of captures
84 3.3.1.2 Catch per unit effort
Gillnet sampling effort in the sampled pools of the Fitzroy River in 2002 to 2009, averaged 243.7 h of 20-m net ± 65.1 (mean ± s.e.) per year. All size classes of P. microdon were captured in the smallest (51 mm) and largest (203 mm) mesh sizes deployed. Catch per unit effort for P. microdon had a general decline between 2003 and 2006 (Fig. 3.4). Catch per unit effort remained relatively constant in late 2004 to
2009, and remained below 10 P. microdon 100 h-1 of 20-m net set, with exception of the early dry in 2005 and 2009 (Table 3.4, Fig. 3.4). Although CPUE of YOY increased in the latter years, CPUE of juveniles >0+ P. microdon, specifically individuals >1800 mm TL, greatly decreased (Fig. 3.3, 3.4). Catch per unit effort of
YOY was observed to significantly decrease between early and late dry seasons (p =
0.041, n = 4) (Table 3.4). A marked decline in older age classes was also observed but was found to not be significant (p = 0.173, n = 4) (Table 3.4). Annual and seasonal patterns in CPUE were found to be similar to the percentage of positive gillnet captures (Fig. 3.4).
Catch per unit effort of YOY in the early (n = 5) or late (n = 4) dry season was not observed to be correlated with total wet season river stage height (p > 0.05).
However, the percentage of YOY to occupy sampled freshwater pools (i.e. CPUE of
YOY in freshwater pools/CPUE of YOY in all pools) in the early dry season was found to be strongly influenced by the total wet season river stage height (r2 = 0.863, p = 0.023, % of YOY in sampled freshwater pools = -7.137 + [0.00455 * total wet season river stage height]). Samples from 2002, 2004 and 2005 were excluded from these analyses due to a lack of 0+ captures and/or fishing effort in the upper or lower pools during the early dry season.
85 In King Sound, fishing effort in the Stokes Bay, SW King Sound and NW
King Sound areas averaged 394.1 h of 20-m net ± 22.1 (mean ± s.e.). Fishing effort in the SE King Sound and Point Torment areas consisted of only 106.7 and 53.9 h of
20-m net sets, respectively. Catch per unit effort was not calculated for P. microdon in King Sound as only one P. microdon was captured.
Table 3.4. Total CPUE (Pristis microdon 100 h-1 of 20-m net set) for YOY and >0+ juvenile
Pristis microdon for the 2002 to 2009 early and late dry seasons, in the Fitzroy River
ED, early dry season; LD, late dry season
Season YOY >0+ 2002 LD 3.2 17.5 2003 ED 4.4 52.4 2003 LD 0 24.5 2004 ED 0 23.0 2004 LD* 0 6.3 2005 ED* 0 12.2 2005 LD - - 2006 ED 1.4 2.3 2006 LD 0.5 1.0 2007 ED 6.0 0.6 2007 LD 0.2 0.2 2008 ED 3.0 1.2 2008 LD 0 1.0 2009 ED 15.4 0 2009 LD* 1.7 1.1
*Estuarine pools were not sampled
86 0.6 100
0.5 80
0.4 60
0.3
40 0.2
20 0.1
Percentage of positive gillnet captures gillnet positive of Percentage
CPUE (number of fish/100 hr of 20-m net set) net 20-m of hr fish/100 of CPUE (number 0.0 0
2002 2003 2004 2005 2006 2007 2008 2009 2010 Date
Fig. 3.4. Total Pristis microdon CPUE (circle) and percentage of seasonal effort with positive Pristis microdon gillnet captures (square) in the early (June to July; black symbol) and late (October to
November; gray symbol) dry season (2002 to 2009); only freshwater pools were sampled in 2005 as well as late 2004 and late 2009.
3.3.1.3 Growth
Mark and recapture data analysis
One hundred and twenty four P. microdon were tagged with Rototags in 2003 to
2009. Of these tagged P. microdon, 31 different P. microdon (excluding visual recaptures and a single recapture that shed its tag) were recaptured between 2005 and
2010. Only five of these P. microdon were recaptured after being at liberty for more than one month (83 to 1246 days), with one being recaptured after two periods of >1 month (Table 3.5). Four of these recaptures were reported by recreational fishers. In most cases, lengths reported by fishers were estimated and thus data from these recaptures should be interpreted with caution. Body length (BL) was measured
87 instead of TL on one occasion, as the sawfish was missing its rostrum. A TL for this
sawfish was calculated using the observed BL to TL relationship: BL = 0.7898 TL –
35.44, which was derived from data from this study and Thorburn et al. (2007). The
resulting growth rates for these recaptured individuals had a mean value of 13 mm
month-1 ± 2.2 (mean ± s.e.) (Table 3.5).
Table 3.5. Growth data of Pristis microdon at liberty for >1 month in the Fitzroy River
Tag #, Rototag number
Capture TL Recapture TL Growth Tag # Capture date (mm) Recapture date (mm) (mm month-1) M1032 30-Jun-08 977 20-Nov-08 1050* 16 F7081 3-Aug-06 1150 26-Oct-06 1170 7 F7092 23-Jun-07 1611 8-Mar-10 1981* 11 F7092 23-Jun-07 1611 20-Nov-10 2273** 16 M1034 17-Sep-09 1865 31-Jul-10 2070 20 F7021 4-Nov-03 2320 10-Aug-06 2540* 7
* Estimated
** TL calculated from BL
Length-frequency data analysis (LFDA)
Length-frequency analyses involved 154 TL measurements pooled into 100-mm size
classes, which ranged from 800 to 2800 mm TL. Only measurements from initial
captures, and recaptures that occurred over one month post initial capture were
included. Known previous capture lengths of 5800 mm (Peverell 2005) and 6560 mm
TL (Compagno and Last 1998) were included in analyses (one of each per sampling
period) to allow for a more appropriate estimate of L. Early dry season months were
pooled, as were late dry season months and sexes due to the limited number of
captures. Years were not pooled as more reasonable curves were produced (i.e.
curves fit the data better) when individual years were analysed.
88 ELEFAN and PROJMAT methods produced k values of 0.106 and 0.141 yr-1, respectively, as well as L values of 5961 and 5825 mm TL, respectively (Table 3.6).
The ELEFAN method produced a better fit to the original data (Fig. 3.5a, b). Hoenig and Choudary Hanumara seasonal growth curves applied to both the ELEFAN and
PROJMAT methods did not greatly improve the score values of these methods (Table
3.6).
Table 3.6. Scores and von Bertalanffy growth function parameters produced from ELEFAN
and PROJMAT methods with non-seasonal and seasonal growth curves
Higher scores represent better fitting curves (PROJMAT and ELEFAN scores are not comparable)
L, to, von Bertalanffy growth parameters; t, age in years; C, extent of seasonal variation in growth;
ts, set point in year in which growth is fastest and slowest
-1 Method Score k (yr ) L (mm) t0 (yr) C ts (yr) PROJMAT (non-seasonal) -2.559 0.141 5825 -0.907 PROJMAT (seasonal) -2.557 0.141 5843 -0.787 0.471 0.254 ELEFAN (non-seasonal) 0.376 0.106 5961 -0.300 ELEFAN (seasonal) 0.376 0.106 5961 -0.270 0.529 0.250
89
Fig. 3.5. Best fit growth curves produced by non-seasonal (a) ELEFAN and (b) PROJMAT methods for juvenile Pristis microdon plotted on original length-frequency data.
3.3.1.4 Mortality and survivorship
Indirect age-specific methods used for estimating M of P. microdon produced instantaneous values between 0.29 and 0.58 yr-1 for 0+ (YOY) P. microdon and 0.11 to 0.20 yr-1 for 5+ P. microdon (Table 3.7, Fig. 3.6). Age-independent methods produced M values ranging between 0.12 and 0.17 yr-1 (Table 3.7). Annual survivorship was estimated to range between 0.56 to 0.75 yr-1 for 0+ P. microdon and
0.83 to 0.89 yr-1 for 5+ P. microdon (Table 3.7). Survivorship calculated from age- independent M values ranged between 0.84 and 0.89 yr-1 (Table 3.7).
90 Table 3.7. Natural mortality (M) and annual survivorship (S) of juvenile Pristis microdon
Peterson and Wroblewski, Lorenzen, and Chen and Watanabe estimate ranges are for ages 0+ to 5+
Model This study Peverell 2008 -1 -1 -1 -1 M (yr ) S (yr ) M (yr ) S (yr ) Pauly (1980) 0.18 0.84 0.15 0.86 Jensen (1996) (1.5) 0.16 0.85 0.12 0.89 Jensen (1996) (1.6) 0.17 0.84 0.13 0.88 Peterson and Wroblewski (1984) 0.29 - 0.13 0.75 - 0.88 - - Lorenzen (1996) 0.29 - 0.11 0.75 - 0.89 - - Chen and Watanabe (1989) - - 0.58 - 0.19 0.56 - 0.83
0.7 Peterson and Wroblewski (1984) Chen and Watanabe (1989) 0.6 Lorenzen (1996)
)
-1 0.5
0.4
0.3
0.2
Natrual mortality (yr mortality Natrual
0.1
0.0
0 1 2 3 4 5 Age (yr)
Fig. 3.6. Age-dependent indirect natural mortality estimates of juvenile Pristis microdon calculated using Peterson and Wrobleski (1984), Chen and Watanabe (1989) and Lorenzen (1996) methods.
3.3.2 Glyphis garricki
Seven G. garricki were captured in creeks and estuaries of King Sound (spanning approximately 130 km of coast line) in June and October 2007 and July 2008 (Table
91 3.8, Fig. 3.7). Total length of captured G. garricki ranged from 1000 to 1365 mm
(Table 3.8). A 1365 mm TL male was found to have fully calcified claspers while a
1090 mm TL male had semi-calcified claspers. Two females of 1088 and 1237 mm
TL were found to have undeveloped ovaries and thus were classified immature.
16°30'S
Robinson River
King Sound 17°0'S
Meda River ) May River Fraser River )
b 17°30'S Western Australia Fitzroy River ¸ 0 5 10 20 30 4040
Kilometers 123°0'E 123°30'E 124°0'E This study ) Previous study
Fig. 3.7. Glyphis garricki capture locations in King Sound in 2007 and 2008 (red dot), and in 2002
to 2004 (blue square) (Thorburn and Morgan 2004).
92 Table 3.8. Glyphis garricki captures in 2007 and 2008 in King Sound
Tag #, Rototag or satellite tag (see Ch. 5) number; SE, South-East King Sound area; SB, Stokes Bay
area; PT, Point Torment area; SW, South-West King Sound area; U, unknown
Tag # Capture date Location Sex TL (mm) SPOTGG1 19-Jun-07 Doctor's Creek (SE) M 1365 M1014 19-Oct-07 Robinson River estuary creek (SB) F 1084 NA 20-Oct-07 North Stokes Bay creek (SB) U 1000 M1009 22-Oct-07 Robinson River estuary (SB) M 1088 NA 22-Oct-07 Robinson River estuary (SB) F 1298 NA 23-Oct-07 May River estuary creek (SB) F 1237 NA* Jun-08 Meda River (SB) - ~1100+ NA* Jun-08 Meda River (SB) - ~1100+ M1068 11-Jul-08 Nearshore channel (SW) M 1090
*Details and pictures submitted by local fishers
Catch per unit effort was 2.3 (n = 5), 1.1 (n = 1) and 0.2 (n = 1) G. garricki
100 h-1 of 20-m net set in the Stokes Bay (October 2007), SE King Sound (June 2007) and SW King Sound areas (July 2008), respectively (see Fig. 3.1). Relative abundance of G. garricki was the third greatest of the captured elasmobranchs in the turbid upper King Sound, which excludes the NW King Sound area (see Fig. 3.1;
Table 3.9).
Water depth, temperature and salinity at the time and space G. garricki were captured at ranged from 1.1 to 6.0 m, 20.2 to 29.5oC, and 32 to 36.8 ppt, respectively.
Two G. garricki were captured on the outgoing tide and five were captured on the incoming tide. Two additional G. garricki were captured (photographed but not measured) on hand lines by local fishers in the intertidal area of the Meda River (~19 km upstream of the river mouth) in June 2008 (Fig. 3.7).
93 Table 3.9. Number of elasmobranchs, by species, captured within the upper King Sound
(excludes the NW King Sound area)
Species Number captured Carcharhinus leucas 13 Pristis clavata 12 Glyphis garricki 7 Carcharhinus cautus 5 Eusphyra blochii 5 Negaprion acutidens 3 Carcharhinus fitzroyensis 2 Pristis microdon 1 Unidentified* 7
*Carcharhinids released prior to identification, likely C. leucas
3.4 Discussion
This study has advanced the understanding of the trends in distribution, population composition, mortality and growth of juvenile P. microdon in the Fitzroy River, and expanded the known distribution of G. garricki within King Sound. Inclusion of data from previous studies (i.e. Thorburn and Morgan 2004; Thorburn et al. 2007;
Thorburn 2008) allowed for the construction of a data set that spanned an eight-year period. The longevity of this study and inclusion of data sourced from previous studies aided in demonstrating CPUE of P. microdon to vary seasonally as well as annually in sampled pools. Additionally, results from this study suggested that growth, mortality and sex ratios change with size of P. microdon and with time.
Finally, this study demonstrated G. garricki to range throughout the upper King
Sound. These results should prove valuable in the development of life history tables of P. microdon and in the understanding of habitat use patterns of these species.
94 3.4.1 Distribution
Pristis microdon
Capture locations of P. microdon in the lower 160 km of the Fitzroy River were similar to those documented by Thorburn et al. (2007). This was expected, as sample sites within the river were similar for both projects. However, this study found that the region inhabited by YOY P. microdon during a particular year was dependent upon the size of the preceding wet season. The increased movement of P. microdon into upstream pools during large wet seasons was likely due to an increase in access to these areas. Observing YOY to move upstream when access was possible could suggest that the freshwater pools provide an ecological advantage to this size class.
As the freshwater pools are environmentally more stable and contain fewer large bodied predators, such as the estuarine crocodile (Crocodylus porosus) and C. leucas, it is reasonable to hypothesise that these upstream pools may aid in increasing survivorship of at least YOY P. microdon.
The sampled area in King Sound was expanded from previous efforts
(Thorburn and Morgan 2004; Thorburn et al. 2007) and encompassed the majority of nearshore waters and tidal-creeks. Even so, the distribution of P. microdon within
King Sound was not expanded. Little is understood about the distribution and habitat use of P. microdon after emigration from its nurseries, due to minimal catches in marine waters (Taniuchi et al. 1991; Peverell 2005; Chidlow 2007; Thorburn et al.
2007; Field et al. 2008; Last and Stevens 2009). However, anecdotal evidence from a local commercial fisher suggests that P. microdon in King Sound is most often captured in open water over sandy substrate. This was supported by the single capture of a sub-adult P. microdon (2580 mm TL male) in open water (~5 km off of
Point Torment) on a submerged sandbar after relatively minimal sampling effort.
95 Only four other P. microdon, all immature (2108, 2280, 2371 and 2440 mm TL), have been recorded in King Sound and were captured within, or in close proximity to the mouths of its tributaries (Thorburn et al. 2007). Although Thorburn et al.’s (2007) captures contradict the fisher’s information, Thorburn’s data could suggest that at least a small percentage and a specific size class of P. microdon (i.e. maturing P. microdon emigrating from rivers) may, at least temporarily, inhabit nearshore water near major tributaries. This is in congruence with findings of P. microdon in
Queensland, Australia (Peverell 2008) and smalltooth sawfish (Pristis pectinata) in
Florida, United States (Simpfendorfer et al. 2010), which found these sawfishes to move from rivers, to tributaries and eventually into deeper marine water with growth.
Glyphis garricki
The known range of G. garricki within King Sound (Thorburn et al. 2004) was expanded upon during this study. Previous records of G. garricki captures within
King Sound have only been from a tidal-creek <20 km north-east of the Derby Jetty and a nearshore channel <7 km south-west of the Derby Jetty. Capture locations now consist of shallow (<6 m) nearshore areas and tidal-creeks extending from the
Robinson River estuary in the Stokes Bay area to near the Fraser River mouth in the
SW King Sound area. This expanse encompasses the entirety of the turbidity maximum zone within King Sound (Wolanksi and Spagnol 2003, Chapter 2). No G. garricki, including those documented from prior studies (Thorburn and Morgan 2004;
Thorburn 2006), were captured outside of this zone in King Sound. Neonate through sub-adult G. garricki have also been captured in similar habitat types throughout its geographic range (i.e. highly turbid and nearshore environments within, or in close proximity to tributaries) (Taniuchi et al. 1991; Larson 2000; Field et al. 2008; Pillans
96 et al. 2009). As noted by Thorburn et al. (2004), the morphology of G. garricki, including its relatively small eyes and high number of sensory pores, makes it well adjusted for environments with low visibility. This would likely provide juvenile G. garricki with an advantage over other competing or predatory species in low visibility environments, which could be reason for this size class to inhabit such areas.
Temperature (20.2 to 29.5°C) and salinities (32 to 36.8 ppt) at capture sites of
G. garricki were consistent with prior findings in King Sound (Thorburn and Morgan
2004; Thorburn 2006). Only minor extensions of minimum water temperature
(20.2°C vs. 21°C) and maximum salinity (36.8 ppt vs. 36.6 ppt) were observed between this study and that of Thorburn (2006).
3.4.2 Population composition
Pristis microdon
Size ranges of male and female P. microdon within the Fitzroy River have been documented to be between 832 and 2770 mm TL and 815 to 2350 mm TL, respectively (Thorburn et al. 2007). This study extended this range, having observed males and females as small as 789 and 763 mm TL, respectively. Observing P. microdon as small as 763 mm TL supports Compagno and Last’s (1998) estimation of size at birth to be as small as 760 mm TL. Similarly, Peverell (2005) observed neonates as small as 720 mm TL and full-term embryos of between 875 and 900 mm
TL (in a 6000 mm TL P. microdon). This evidence implies a comparable pupping size of P. microdon in the Gulf of Carpentaria and Western Australia. Having not observed females greater than 2770 mm TL and males greater than 2350 mm TL after eight years of sampling the Fitzroy River, supports the hypothesis that females and
97 males emigrate from the river into estuarine and marine waters at sizes less than 2800 and 2400 mm TL, respectively (Thorburn et al. 2007).
Sex ratio of P. microdon in the Fitzroy River appeared female dominant, with a mean sex ratio of 1: 1.5 ± 0.30 (mean ± s.e.). However, upon closer examination, yearly sex ratios appeared to be biased by the presence of P. microdon >2400 mm TL
(P. microdon >2400 mm TL in the river were all female). Sex ratios of smaller P. microdon varied annually but in general were approximately 1: 1. Sex ratios became female dominant when P. microdon >2400 mm TL were present. The transition in sex ratios could suggest that when a cohort reaches around 2400 mm TL, that the number of males decreases or the number of females increases. A decrease in males could be evidence that males emigrate from the Fitzroy River at an earlier age than females. Alternatively, an increase in the number of females may result from the introduction of additional females from neighboring rivers. In either situation, males appear to inhabit riverine environments for a shorter period than females. Using estimated ages of the different size classes, which are based off this study as well as
Thorburn et al. (2007) and Peverell (2008), male and female P. microdon appear to begin to transition to marine waters three to four and three to five years, respectively, post parturition. Thorburn et al. (2007) also hypothesised that males may emigrate before females, suggesting this may be the cause of the difference in maximum size between the sexes. However, he also stated that the difference in sizes may be a result of different rates of growth. Sex-based differences in timing (months to years) of nursery area emigration may increase female survivorship and therefore chance of reproduction, and has been observed in other species (Jonsson et al. 1990; Oliveira
1999; Heupel 2007).
98 Glyphis garricki
Size ranges of male and female G. garricki captured within King Sound was similar to that documented by Thorburn and Morgan (2004) (904 to 1418 mm TL). However, the 1: 1 sex ratio observed in this study varied from previous findings (Thorburn and
Morgan 2004; Thorburn 2006). This is likely due to the small sample sizes observed in this and previous studies. Compilation of all known G. garricki within King Sound resulted in a sex ratio of 1: 0.6 (male: female).
Size at maturity appeared to occur by at least 1365 mm TL for males. This was in agreement with previous findings that suggested males mature between 1350 and 1420 mm TL (Thorburn and Morgan 2004; Pillans et al. 2009). Maturity of females has been suggested to occur by 1770 mm TL (Pillans et al. 2009) and to not occur at <1350 mm TL (Thorburn 2006). This was supported by the observation of only immature females that were all <1300 mm TL. Juveniles appear to be the dominant G. garricki life-stage observed in rivers and turbid nearshore waters in northern Australia (Taniuchi et al. 1991; Field et al. 2008; Pillans et al. 2009).
However, it is also likely that this is a product of size-selectivity of sampling methods.
Using a 178-mm mesh, as was primarily used here, McAuley et al. (2007) found the maximum selectivity of this mesh size for the less stout sandbar shark (Carcharhinus plumbeus) to be between 950 and 1240 mm FL. This is a similar size range to that of the G. garricki captured in this study. Only the largest (1365 mm TL) G. garricki captured was done so in a 206-mm mesh gillnet. Incorporating varying sampling methods including nets with mesh >206 mm, may provide a better estimate of the size range of G. garricki in King Sound.
99 3.4.3 Population life history of Pristis microdon
3.4.3.1 Growth
Observed growth of P. microdon from mark and recapture methods was limited to a few individuals, but was similar to growth rates of juveniles observed by other studies
(e.g. Thorburn et al. 2007, Peverell 2008). However, the majority of estimates in this study were towards the lower ranges described by Thorburn et al. (2007) and Peverell
(2008). Having such a limited dataset it was difficult to discern the cause of such a large range in growth rates. However measurement error/estimation was likely to be largely responsible for the variance, as recapture measurements were often estimated by recreational fishers.
In the light of uncertainty in the accuracy of recapture length estimates, length-frequency analysis was also employed (lethal vertebral analysis was not used, due to the threatened status of the species). This method was proven to be useful with juvenile P. pectinata (Simpfendorfer et al. 2008b), and in this study the method produced estimates that closely fit the data. Of the two methods used in the length frequency analyses, the ELEFAN method gave a visually better fit to the data. While
L estimates were similar to back calculated data produced from vertebral analysis of
Peverell (2008), the k value of 0.106 yr-1 was greater than previous estimates of 0.066 yr-1 (Tanaka 1991) and 0.08 yr-1 (Peverell 2005, 2008). This was expected as
Tanaka’s k value has been shown to greatly underestimate growth of P. microdon
(Thorburn et al. 2007; Peverell 2008). Peverell (2008) also reported his k value to likely underestimate growth (after observations of faster growth from mark and recapture methods). In this study, the length-frequency analysis was dominated by fast growing juveniles and therefore best represents growth of juvenile P. microdon
(800 to 2800 mm TL). While growth of juvenile P. microdon estimated by this study
100 was greater than previous findings, it was slower than that described for P. pectinata
(k = 0.140 yr-1) (Simpfendorfer et al. 2008b).
Inclusion of seasonal growth curves did not substantially improve the fit of the curve to the data, suggesting continual annual growth. This result was unexpected due to the significant changes in seasonal temperatures (Chapter 2) and strong correlation between temperature and growth in fishes (Pauly 1980; Carrier and Luer
1990; Natanson 1993). Lack of evidence supporting seasonal growth may have been a result of the lack of wet season data, and/or due to the pooling of months into only two points per year in the analyses. Simpfendorfer et al. (2008) observed juvenile P. pectinata in subtropical environments to exhibit weak seasonal growth by pooling years and separating the year into multiple sampling periods. Unfortunately, pooling years and separating months in attempts of detecting seasonal growth produced poor fits to the data of this study and therefore were not included in final analyses. It is believed that this poor fit may have been a result of the relatively few 1+ P. microdon captured in this study, as this age class was omitted by growth curves years were pooled.
3.4.3.2 Mortality
Indirect natural mortality estimates of P. microdon produced from the age- independent methods were similar and ranged between 0.12 and 0.18 yr-1.
Simpfendorfer et al. (2000) reported similar values for largetooth sawfish (Pristis perotteti) (M = 0.069 to 0.155 yr-1) and P. pectinata (M = 0.059 to 0.139 yr-1) after using similar methods. This study also found the age-specific Lorenzen (1996) and
Peterson and Wroblewski (1984) methods to produce similar results (0+ to 5+ P. microdon = 0.29 to 0.11 yr-1) to those observed for P. perotteti (0+ to 5+ P. perotteti =
101 0.27 to 0.14 yr-1) (Simpfendorfer 2000). Observing similar values of M for P. microdon and P. perotteti would be reasonable as these species occupy similar habitats and have similar life histories (Thorson 1982; Simpfendorfer 2000;
Simpfendorfer et al. 2010). Only the Chen and Watanabe (1989) method produced substantially different (greater) estimates of M in comparison to other methods within this study, as was also observed by Simpfendorfer (2000). Unfortunately, it is not possible to ascertain which method is more accurate without direct mortality estimates. In theory, the M estimate for YOY produced by the Chen and Watanabe
(1989) method may be the most accurate. In a study comparing direct and indirect estimates of M, Heupel and Simpfendorfer (2002) found indirect methods to underestimate M. In light of this uncertainty, it is recommended to use the most extreme values of M (i.e. M produced by the Chen and Watanabe method) to err on the side of caution for this critically endangered species, until direct measurement of
M is available.
3.4.4 Catch per unit effort
Pristis microdon
Catch per unit effort of P. microdon was observed to fluctuate between years in the sampled pools of the Fitzroy River, with a general decrease between 2002 and 2006.
This decrease was attributed to a drop in captures of P. microdon >1850 mm TL. The decrease in number of this specific size class (relative to years prior) may have been the result of a previous small year-class growing-up (McGlennon et al. 2000;
Anderson & Rose 2001). Indeed, CPUE of YOY P. microdon in the early 2004 to
2006 dry seasons was 0 to 23 % of the mean CPUE in all other years. The small number of YOY observed within these years may have been the result of a change in
102 the survivorship of YOY cohorts (Gruber et al. 2001; Heupel and Simpfendorfer
2002), an artifact of sampling methods (e.g. sample sites may have been spaced too far apart and allowed for a portion of YOY to go undetected in certain years) and/or the result of fluctuations in the size and health of the breeding stock (Kinney and
Simpfendorfer 2008; Dulvy and Forrest 2010). However, the effect of the latter argument is unlikely to have been observed on an annual scale.
Alternatively, or additionally, the decline in CPUE may have been due to an abnormal peak in CPUE of >1800 mm TL P. microdon in the early 2000s, resulting from a highly successful year class in a previous year or years. Interestingly, the 1999 and 2000 wet seasons each produced a river discharge and stage height total greater than all other subsequent years. It was during 1999 and 2000, that the P. microdon, which composed the abnormal peak, would have been pupped. The hypothesis that survivorship is positively related to movement upstream and thus total wet season stage height, would help explain why such a relatively large number of P. microdon was present in the river during and after the two subsequent large wet seasons.
Catch per unit effort also varied seasonally. Seasonal fluctuations in CPUE suggest that the number of P. microdon within a pool greatly changes through the year in the Fitzroy River. This change appears to be size dependent. Young of the year demonstrated a substantially greater decrease (mean = 90 %) in CPUE between the early and late dry seasons than their older conspecifics (mean = 51 %). Reasoning for this difference is most likely due to the relatively high rate of mortality of YOY.
This is supported by the age-specific estimates of M, which demonstrated that annual
M is greater for YOY, and accounts for more than half of the decline in CPUE of
YOY and >0+ P. microdon in the Fitzroy River. However, YOY also have the increased ability, in comparison to larger juveniles, to move through shallow runs
103 between pools and out of the study site (Chapter 4). If YOY were moving out of the study sites on a more frequent basis than larger individuals, it would result in the calculation of a greater decrease in CPUE of YOY. Fishing mortality from recreational fishers may also contribute to the seasonal changes in CPUE, but this has not been documented within the Fitzroy River.
Seasonal differences in CPUE of certain fishes has been suggested to be the result of a change in catch efficiency of fishing methods brought on by a change in fish behavior (Pope and Willis 1996). However, this was not believed to have caused the observed seasonal shifts in CPUE of P. microdon. The capture of P. microdon via gillnets is likely to only increase in the late dry season, rather than decrease, as P. microdon movement increases with water temperature (Peverell pers. comm.).
Glyphis garricki
Catch per unit effort of G. garricki was documented to range between 0.2 and 2.3 fish
100-1 h of 20-m net set in King Sound. However, it is difficult to conclude about the patterns in G. garricki CPUE due to small sample sizes and lack of repetition. Even so, it is should be noted that the greatest CPUE of G. garricki occurred within the
Stokes Bay area, which is fed by three of the four largest rivers (i.e. Robinson, May and Meda Rivers) entering King Sound. As G. garricki have previously been shown to inhabit rivers and river mouth areas as neonates and juveniles (Tanaka 1991; Field et al. 2008; Pillans et al. 2009), the confluence of three large rivers in the Stokes Bay area may provide this species with suitable habitat. Interestingly, G. garricki have never been recorded in the Fitzroy River, unlike in the Meda River.
The status and size of the Australian G. garricki population is unknown but has been suggested to be small due to its apparent rarity (Pogonoski and Pollard
104 2003). During this study however, G. garricki was the third most abundant of nine elasmobranchs captured in the highly turbid upper King Sound. Glyphis garricki captures were only outnumbered by P. clavata (n = 12), also listed as Critically
Endangered on the IUCN Red List of Threatened Species (Cook et al. 2006), and C. leucas (n > 13). However, the catchability of sawfishes is likely to be greater because of the morphology of their rostrum. Prior sampling in the SE King Sound area and estuarine pools of the Fitzroy River by Thorburn et al. (2004), also observed G. garricki to be one of the more abundant (third of eight) elasmobranch species captured (nG. garricki = 6 of ntotal = 46). From this evidence it would appear that G. garricki is one of the more common elasmobranch species, which are susceptible to gillnets, to inhabit the highly turbid nearshore waters of King Sound. This could suggest that G. garricki are not as rare as currently thought and/or that G. garricki, C. clavata and C. leucas have a greater affinity for this habitat type than other elasmobranchs. However, further sampling is needed to provide statistically robust evidence of this conclusion.
3.4.5 Conclusion
Data from this study suggests that P. microdon inhabit the Fitzroy River for between three to five years. During this period growth is rapid, similar to P. perotteti and P. pectinata (Simpfendorfer et al. 2008b), with individuals approximately tripling in length. Distribution of P. microdon in the Fitzroy River was dependent upon wet season size/stage height, with a greater percentage of YOY inhabiting upstream pools in years with large wet seasons. The composition of P. microdon in the study area varied temporally at a seasonal and annual scale. While seasonal changes in relative abundance appeared to be most influenced by natural mortality, reasoning for
105 variations in annual CPUE was unclear. Although annual CPUE was hypothesised to be influenced by river stage height and/or discharge, no evidence was found to support this. However, it is the view of the author that this relationship should be further explored in future research as sampling methods may have biased results.
Together the findings of this study demonstrate that the composition and distribution of juvenile P. microdon in the Fitzroy River is not static, due to the influence of multiple abiotic and biotic variables. As such, conservation strategies will need to be flexible and take into account potential size and sex-based differences of P. microdon when assessing how natural and anthropogenic environmental changes will impact this species.
Results from this study also suggested G. garricki to be more abundant in the turbid nearshore water of the upper King Sound than most other elasmobranch species captured in this study. However, these results were based on a small number of captures and needs to be expanded upon with further sampling. Results also suggested that at least juvenile G. garricki may be limited to highly turbid nearshore waters, whether it is for predator avoidance or increased capture of prey. Although results from this study were limited, the acquired information has provided insight into the specifics of the habitat in which juvenile G. garricki occur. By increasing our knowledge of G. garricki habitat, researchers can better explore why and how G. garricki use these areas and can better predict areas in which G. garricki may be found. Through the combination of simple visual analysis of satellite imagery of northern Australia with known characteristics of juvenile G. garricki habitat, it is evident that potential habitat for this species is greatly limited.
106 Chapter 4: Depth use and inter-pool movement of
juvenile freshwater sawfish Pristis microdon in the
Fitzroy River, Western Australia
4.1 Introduction
Elasmobranch nursery areas are important habitats that allow for increased survival of juveniles (Beck et al. 2001; Heupel et al. 2007), due to advantageous characteristics such as relatively high levels of prey and/or reduced predation levels (Castro 1993;
Beck et al. 2001). In northern Australia, the freshwater sawfish (Pristis microdon) has been shown to use riverine environments as nursery areas for their first few years of life, before emigrating into marine waters (Thorburn et al. 2007). Many of the large rivers in northern Australia, where this species has been observed, have been and/or are being targeted for development of dams, mining and/or water abstraction
(Wolanski et al. 2001; Doupé and Pettit 2002; Lindsay and Commander 2006;
Thorburn and Morgan 2006; CSIRO 2009). Already, installation of a dam in the Ord
River has dramatically transformed its flow regime (Wolanski et al. 2001; Doupé and
Pettit 2002). However, it is uncertain how any environmental alterations would affect
P. microdon either directly or indirectly, via changes to prey species, due to the paucity of information pertaining to P. microdon habitat use. Observations of other elasmobranchs within riverine and estuarine settings, whose habitats were altered by construction and mining projects, urban run-off as well as water regulating dams, have shown a decrease in ontogenetic resource partitioning (Carlisle and Starr 2009), emigration of juveniles into less favourable environments with greater risks of predation (Simpfendorfer et al. 2005; Carlisle and Starr 2009) and/or reductions in
107 survival rates (Jennings et al. 2008). It is reasonable to suggest that similar results may occur with P. microdon when individuals are exposed to similar circumstances.
Observing the lack of information regarding habitat use of P. microdon, this study aimed at exploring how changes in abiotic and biotic variables influence habitat use of juvenile P. microdon. It was hypothesised that the habitat use patterns of juvenile
P. microdon would be influenced by the significant changes in salinity, water temperature and turbidity/light intensity that occur in the Fitzroy River (Chapter 2), as has been documented with other elasmobranchs (Cartamill et al. 2003; Hopkins and
Cech 2003; Dewar et al. 2008; Heupel and Simpfendorfer 2008), including the smalltooth sawfish (Pristis pectinata) (Simpfendorfer et al. 2011). Additionally, it was hypothesised that river flow and depth would influence the behaviour of P. microdon, as habitat use is subject to habitat accessibility and availability. Passive acoustic monitoring of juvenile P. microdon in the Fitzroy River, Western Australia, was used in association with long-term environmental monitoring to test the stated hypotheses.
4.2 Material & Methods
4.2.1 Study site
Depth use and inter-pool movements of juvenile P. microdon were documented in the lower 160 km of the Fitzroy River, Western Australia, between June 2007 and
November 2009 (see Fig. 4.1). The physical environment within this area greatly fluctuates through the year between wet and dry seasons (see Chapter 2). River stage height decreases in the dry season due to minimal precipitation and increased evaporation. This decrease in river level results in the formation of pools that are
108 connected by shallow runs (generally <0.5 m in depth), if at all. Pool lengths vary during the dry season, with the larger pools spanning up to several kilometres in length. Increased precipitation in the wet season causes a substantial rise in stage height and often floods the river banks.
A total of nine pools were monitored by acoustic receivers for the presence of
P. microdon tagged with acoustic transmitters (Fig. 4.1). These included the tidally influenced estuarine pools: Lower Pelican Pool (the river mouth), Milli Milli (5 km upstream of the river mouth), Snag Pool (6.2 km upstream of the river mouth),
Telegraph Pool (11.6 km upstream of the river mouth) and Langi Crossing (14.9 km upstream of the river mouth), as well as freshwater pools: Myroodah Crossing (123 km upstream of the river mouth), Camballin Pool (156.5 km upstream of the river mouth), Lower Barrage Pool (160.7 km upstream of the river mouth) and Upper
Barrage Pool (161.5 km upstream of the river mouth) (Fig. 4.1). To note, names for
Lower Pelican Pool, Camballin Pool, Lower Barrage Pool and Upper Barrage Pool were given by the author as no formally documented names could be found for these pools. The estuarine pools were approximately 0.1 to 0.2 km in width and approximately 1 to 2.5 km in length, with exception of Lower Pelican Pool (0.5 km in length), in the dry season. Depths of the estuarine pools increased with distance upstream from the river mouth, from 2.0 m at Lower Pelican Pool to 3.5 m at Langi
Crossing (see Chapter 2). Monitored freshwater pools were approximately 0.06 km in width and 1 to 2 km in length in the dry season (see Chapter 2). Depth of the freshwater pools was up to 4.2 m, with occasional deeper holes of over 6 m along the meanders of the river (see Chapter 2). Salinity, diel temperature fluctuation and turbidity decreased with transition from estuarine to freshwater pools, while
109 vegetation diversity and concentration greatly increased (Storey et al. 2001; see
Chapter 2).
Fig. 4.1. Acoustic monitoring stations (red dots) in the Fitzroy River, Western Australia.
Crs. = Crossing, Lwr = Lower, Pl = Pool, Upr = Upper.
4.2.2 Acoustic array and range testing
Six Vemco VR2W acoustic receivers (VEMCO Division, AMIRIX Systems, Inc.,
Halifax, Nova Scotia, Canada) were installed in Lower Pelican Pool, Milli Milli, Snag
Pool, Langi Crossing, Myroodah Crossing (receiver referred to as “Myroodah
Crossing 1”) and Camballin Pool (see Fig. 4.1) in June 2007. The Lower Pelican Pool receiver was moved to Telegraph Pool (Fig. 4.1) because of the shifting river channel in Lower Pelican Pool in July 2008. Receivers were installed in Lower (~0.4 km
110 downstream of the Camballin Barrage) and Upper Barrage Pool (~0.5 km upstream of the Camballin Barrage), as well as in the upper portion of Myroodah Crossing
(receiver referred to as “Myroodah Crossing 2”; located ~1 km upstream of the
Myroodah Crossing 1 receiver) (see Fig. 4.1) in December 2008. Pool locations were chosen based on a combination of accessibility, known P. microdon habitat and experimental array design. The receiver array was designed as a series of gates that would monitor the passing of tagged P. microdon up and down the river. By spacing out groups of receivers (receiver groups consisted of a small number of receivers placed within relatively close proximity to one another in a region of the river), the array design could monitor short and long-distance movements between pools with a small number of receivers. Receiver locations in each pool were chosen based on the depth, available anchoring points and lack of obstructions to acoustic signals (e.g. trees and river bends).
Acoustic receivers were secured to moorings within their respective pools.
Moorings were designed to allow for retrieval of receivers from the water’s surface, as diving was not possible. Mooring designs consisted of a permanent primary anchor
(i.e. concrete block, piling or snag/tree) connected to a secondary smaller sand anchor via a 4-m galvanized chain. The secondary anchor was in place to reduce movement/noise and entanglement of the 4-m chain and 3-m rope (described below), while allowing retrieval of acoustic receivers using the total 7 m of chain and rope. A
3-m rope (25 mm diameter) joined the sand anchor to surface and subsurface aluminium floats. Floats were set 1 m apart and used to compensate for rising and falling river levels. Aluminium floats were used to reduce damage to the floats by crocodiles. Aluminium floats were switched with closed-cell extruded polystyrene foam floats in 2008, after a number of aluminium floats sank after being vandalised.
111 Following the design of Clements et al. (2005), a 1-m rope (25 mm diameter) was connected 1 m above the anchor and 0.5 m below the subsurface float, creating a loop that ran parallel to the main rope. Receivers were attached upright (i.e. the hydrophone was facing up) to the loop using five large cable ties. The purpose of the loop was to maintain receivers in an upright position during strong flows to maximise the range of detections. Receivers were moved in 2009, to approximately 1 m below the surface, with the hydrophone facing downward, to allow for easier retrieval.
Range tests were conducted on deployed receivers using Vemco V13-1L and
V16-5H acoustic ‘test tags’ in June and/or October 2007. Range tests were conducted to test the distance in which each tag type could be detected by VR2W acoustic receivers within the respective pools. Test tags were set to continuously transmit in three to six second intervals. Range tests were conducted in all pools with exception of Telegraph Pool as well as Lower and Upper Barrage Pools, as these were installed after the battery life of the test tags had expired. Range tests consisted of mooring test tags 1 m above the bottom of the river for a five minute period at intervals of 0.05 km, upstream and downstream of the receiver, for distances of between 0.25 and 1.20 km.
The distance of range tests were dictated by pool depth and shape, and were confined to relatively deep (>0.5 m in depth) and straight portions of the river. An additional range test was conducted in Langi Crossing with V13 tags to explore how a silicone coating on acoustic tags would affect transmission strength, as a silicone coating was used for acoustic tag attachment (see below). Research vessels were removed from the area and the engine stopped during range tests to prevent interference with acoustic transmissions.
112 4.2.3 Tagging
Pristis microdon was captured in the river using 20-m gillnets composed of 76, 102,
152 and 203-mm stretched monofilament as well as hook and line methods in 2007 to
2009. These sawfish were moved to the shoreline ensuring that their spiracles, gills and mouth remained submerged, and were inverted on to their dorsal surface, which caused them to enter tonic-immobility. Sawfish were then measured, sexed and tagged (see below). Small P. microdon <1200 mm total length (TL) were occasionally measured and tagged on a research vessel, which reduced overall handling time.
A total of 27 acoustic tags were deployed on 26 juvenile P. microdon (one individual was tagged twice as its original acoustic tag was shed during recapture) between June 2007 and May 2009. Tagged P. microdon ranged between 842 and
2210 mm TL (Table 4.1). Deployed acoustic tags consisted of Vemco V9, V13 and
V16 coded acoustic tags. Tags transmitted at 69 kHz at a random interval that ranged between 45 and 90 s. Power levels of tags also varied. Low power V13-1L tags were most commonly deployed. However, high power V16-5H tags were also initially deployed in case the strength of V13 tags was inadequate. The V9-2H and V13-1H tags, which were deployed in 2009, were donated from previous studies and determined to be adequate for this research. Differences in the powers of the tags may have altered the upstream and downstream distance that P. microdon was detected. However, all deployed tags were unlikely to go undetected if an individual passed a receiver, due to the narrow width of the river (see “Results”). The majority of tags (all but tags 21-23, 25 and 26) contained pressure sensors, which allowed for the monitoring of tag depth. The V9 and V13 tags were secured to Rototags (Dalton
Supplies, New South Wales, Australia) using small cable ties. Cable ties were coated
113 in marine-grade silicone to further secure the tags and prevent any potential abrasion from the cable ties on the fish. Modified Rototags were then fitted to the first dorsal fin in the same manner as unmodified Rototags (Heupel et al. 1998; see Chapters 3,
5). The V16 tags were enclosed in shark casings (a protective covering with anterior and posterior attachment holes) and fitted to the first dorsal fin. A prefabricated drill template and support were used in conjunction with an implant-grade stainless steel drill bit to create two 5-mm holes within the dorsal fin. Hex-head silicone bolts (4 mm diameter) were passed through the dorsal fin and tag attachment holes, and secured via stainless steel washers and nuts.
Table 4.1. Acoustic tag deployment information for 2007 to 2009
Tag #, Acoustic tag number
TL Tag # Tag model Rototag # (mm) Capture date Capture location 1 V16TP-5H F9835 1555 22-Jun-07 Snag Pool 2 V16TP-5H F7092 1611 23-Jun-07 Snag Pool 3 V16TP-5H M1013 1580 18-Jul-07 Snag Pool 4 V13TP-1L M1007 842 18-Jul-07 Snag Pool 5 V13TP-1L M1006 990 19-Jul-07 Snag Pool 6 V13TP-1L M1003 981 19-Jul-07 Snag Pool 7 V13TP-1L M1005 854 19-Jul-07 Snag Pool 8 V13TP-1L M1004 1111 1-Nov-07 Camballin Pool 9 V16TP-5H NA 1576 13-Nov-07 Langi Crossing 10 V13TP-1L M1002 1897 24-Jun-08 Myroodah Crossing 11 V16TP-5H NA 1429 29-Jun-08 Snag Pool 12 V13TP-1L M1038 1423 29-Jun-08 Snag Pool 13 V13TP-1L M1022 918 29-Jun-08 Snag Pool 14 V13TP-1L M1023 959 30-Jun-08 Snag Pool 15 V16TP-5H NA 1934 22-Jul-08 Snag Pool 16 V13TP-1L M1048 2142 27-Oct-08 Myroodah Crossing 17 V13TP-1L M1045 / 2210 28-Oct-08 / Camballin Pool M1050 20-Nov-08 18 V13TP-1L M1051 1883 31-Oct-08 Snag Pool 19 V13TP-1L M1049 2090 31-Oct-08 Snag Pool 20 V13TP-1L M1047 2102 31-Oct-08 Snag Pool 21 V13-1H M1077 965 12-May-09 Snag Pool 22 V9-2H M1076 952 13-May-09 Camballin Pool 23 V9-2H M1078 1027 13-May-09 Camballin Pool 24 V13TP-1L M1066 1013 14-May-09 Lower Barrage Pool 25 V13-1H M1079 872 14-May-09 Lower Barrage Pool 26 V13-1H M1081 911 14-May-09 Lower Barrage Pool
114 4.2.4 Environmental conditions and prey abundance
Water depth, salinity, temperature and visibility/turbidity were recorded manually at each receiver download and during range tests (see Chapter 2) in 2007 to 2009. In addition, HOBO Water Temp Pro V2 Data Loggers (Onset Computer Corporation,
Bourne, MA, USA) were installed on moorings to monitor pool temperatures between
2007 and 2009. HOBO Pendant Loggers (Onset Computer Corporation, Bourne, MA,
USA) were also installed on moorings to measure light intensity and temperature during the 2009 wet season, and in short intervals of time in 2008. Pendant Loggers were deployed in intervals of 0.5 m on moorings, starting at a depth of 0.5 m (see
Chapter 2). River stage height and discharge data, solar data as well as tide information for the Derby area were obtained from the Department of Water,
Government of Western Australia, the United States Naval Observatory Astronomical
Application Department (http://aa.usno.navy.mil/data) and the Western Australian
Department of Planning and Infrastructure, respectively.
Relative prey abundance for sandbar habitats (areas of documented sawfish feeding) was measured in Myroodah Crossing and Camballin Pool. Estuarine pools were not sampled for prey due to abundance of debris and potential danger from crocodiles. Sampling was conducted using a 26-m seine net consisting of two 8-m wings (6-mm mesh) and a 10-m pocket (3-mm mesh), which was capable of fishing depths of 1.5 m. Three samples were taken at each pool at 10:00 h and again at 20:00 h (i.e. six samples were taken per day) in June and October of 2007 and 2008.
115 4.2.5 Data analysis
Statistical analyses were conducted using SPSS V15 (SPSS, Chicago, Illinois, USA) or SigmaPlot 11.0 (SPSS, Chicago, Illinois, USA). Maps were produced with ArcGIS
9.3.1 (ESRI, Redlands, California, USA).
Detection range of VR2W acoustic receivers was calculated by comparing the number of transmissions recorded during range tests, with the number of transmissions expected to occur at each 50-m interval (expected = test duration x tag transmission rate). Receivers were determined to be reliable in detecting transmissions at a given distance if 95 % of expected transmissions were observed at that distance. A Student’s paired t-test was run to test for differences in the range in which silicone and non-silicone coated V13 tags were detected.
Differences between the depth use of P. microdon size classes (pooled into
<1000 mm, 1400 to 1650 mm and >1850 mm TL size classes), seasons and hours of the day were explored with the assistance of tag depth sensor data. Size classes were chosen based on their approximate ages and roughly equated to: young of the year
(YOY), 1+ to 2+ and >2+ P. microdon (Thorburn et al. 2007; Peverell 2008). Only depth readings from individuals present for every hour of the day, with approximately
30 detections (i.e. recorded transmissions) per hour of the day were included in statistical analyses and figures concerning depth. This was done to decrease potential bias from individuals present only during certain hours. Depth profiles of the different size classes were visually, rather than statistically, analysed due to the extreme differences in their distributions. However, differences of depths occupied by individuals in the dry and wet seasons were statistically analysed by running
Mann-Whitney rank sum tests for each individual. Only data from the two P. microdon inhabiting the freshwater pools was included in this analysis, due to
116 minimal wet season data of P. microdon in the estuarine pools. Differences in depth use between hours of the day were also analysed by a repeated measure one-way
ANOVA, as depth use between hours was not independent. Random sub-samples of depths recorded from each hour were selected to create equal sample sizes in order to increase the robustness of analyses. The strengths of the relationships between water depth of P. microdon and light intensity as well as water temperature were examined with a Spearman rank order correlation.
Catch per unit effort (CPUE) of species P. microdon is known to prey on, including the lesser salmon catfish (Neoarius graeffei), bony bream (Nematalosa erebi) and cherabin (Macrobrachium rosenbergii) (Thorburn 2006; Thorburn et al.
2007) as well as other potential prey species captured in seining efforts, were calculated. A two group Mann-Whitney test was run to test for differences between day and night CPUE for the more abundant species. Catch per unit effort of replicates, pools, seasons and years were pooled.
4.3 Results
4.3.1 Acoustic array and range testing
Efforts to download receivers occurred at least twice a year, although inaccessibility and mechanical failure prevented several downloads and an incontinuity in receiver coverage (Fig. 4.2). Inaccessibility was due to receiver burial by shifting sandbars at
Lower Pelican Pool in the 2007 dry season, and by a landslide in Camballin Pool during the 2008 wet season. Missing receivers at Snag Pool and Langi Crossing in
2009, were also considered inaccessible. Buried receivers were recovered and the
Lower Pelican Pool receiver was moved to Telegraph Pool. Electrical errors occurred
117 in four receivers between November 2007 and 2008, as well as in one in Telegraph
Pool in 2009. Non-functional receivers were replaced when possible.
Range tests demonstrated receiver detection radiuses for 95 % of expected
V16-5H and V13-1L transmissions to be up to at least 0.2 (up to 0.45 km in larger pools) and 0.15 km (up to 0.35 km in larger pools), respectively, in all pools. This distance was greater than the width of the river at each receiver. The number of detections rapidly decreased beyond these distances, with 0 % of expected transmissions being detected at an additional 0.10 to 0.15 km. No significant difference was observed between V13 tags with and without a silicone coating (p =
0.383, d.f. = 14).
4.3.2 Acoustic tag performance
Twenty-one of 26 tagged P. microdon were recorded on at least one receiver between
June 2007 and November 2009 (Table 4.2, Fig. 4.2). Two undetected tags were deployed in isolated freshwater pools in which receivers malfunctioned (see Table
4.2, Fig. 4.2). Three undetected tags were deployed in Snag Pool when only a single receiver at Milli Milli or no receivers within the region were functioning (see Table
4.2, Fig. 4.2). Mean time between deployment and last recorded transmission (i.e. days at liberty) for V16 tags was 77.2 days ± 23.3 (mean ± s.e.), and ranged between
27 and 139 days for V16 tags. Mean time at liberty for V9 and V13 tags was 73.6 days ± 19.4 (mean ± s.e.), and ranged between 12 and 313 days. V16 tags were detected by a receiver at least once per day for between 11 and 64 days (Table 4.2).
V9 and V13 tags were detected by a receiver at least once per day for between 3 and
61 days (Table 4.2).
118 Table 4.2. Detection details of acoustic tags deployed in 2007 to 2009
Tag #, Acoustic tag number; LD, Last detection
Days Tag Tag TL Tagging Days LD location LD date at # model (mm) Location detected liberty 1 V16TP-5H 1555 Snag Pl Snag Pl 17-Aug-07 56 19 2 V16TP-5H 1611 Snag Pl Snag Pl 30-Jul-07 37 19 3 V16TP-5H 1580 Snag Pl Snag Pl 14-Aug-07 27 14 4 V13TP-1L 842 Snag Pl Langi Crs. 30-Sep-07 74 17 5 V13TP-1L 990 Snag Pl Snag Pl 2-Oct-07 75 33 6 V13TP-1L 981 Snag Pl Snag Pl 13-Aug-07 25 15 7 V13TP-1L 854 Snag Pl Snag Pl 21-Aug-07 33 22 8 V13TP-1L 1111 Camballin Pl - - - - 9 V16TP-5H 1576 Langi Crs. Langi Crs. 31-Mar-08 139 11 10 V13TP-1L 1897 Myroodah Crs. - - - - 11 V16TP-5H 1429 Snag Pl Langi Crs. 3-Nov-08 127 64 12 V13TP-1L 1423 Snag Pl - - - - 13 V13TP-1L 918 Snag Pl Telegraph Pl 30-Sep-08 93 47 14 V13TP-1L 959 Snag Pl Telegraph Pl 30-Aug-08 61 27 15 V16TP-5H 1934 Snag Pl - - - - 16 V13TP-1L 2142 Myroodah Crs. Milli Milli 13-May-09 198 54 17 V13TP-1L 2210 Camballin Pl Myr. Crs. 12-Jan-09 76 61 18 V13TP-1L 1883 Snag Pl Milli Milli 9-Sep-09 313 26 19 V13TP-1L 2090 Snag Pl Milli Milli 15-Dec-08 45 13 20 V13TP-1L 2102 Snag Pl Milli Milli 15-Nov-08 15 3 21 V13-1H 965 Snag Pl - - - - 22 V9-2H 952 Camballin Pl Camballin Pl 2-Jul-09 50 20 23 V9-2H 1027 Camballin Pl Camballin Pl 30-Jun-09 48 35 24 V13TP-1L 1013 Lwr Barrage Pl Lwr Barrage Pl 1-Jun-09 18 6 25 V13-1H 872 Lwr Barrage Pl Lwr Barrage Pl 25-Jun-09 42 34 26 V13-1H 911 Lwr Barrage Pl Lwr Barrage Pl 26-May-09 12 3
119
rossing,
are representedare by
pper Barrage Pool.
Activity ofActivity (top) acoustic tags and (bottom)hydrophones in the respectivepools. Tag deployments
Fig. 4.2. trianglesand detected aretransmissions represented 1: by circles. Milli Milli, 2:Snag Pool, 3:Telegraph Pool,4:Langi C 5:Myroodah Crossing 6: 1, Myroodah Crossing 7: 2, Camballin Pool, 8:Lower Barrage Pool, 9:U
120 4.3.3 Depth use
Depth profiles of P. microdon <1000 mm TL in the estuarine pools were heavily skewed to the right (with exception of tag 13), with individuals spending 90 to 95 % of their time between 0 ± 0.00 and 1.3 m ± 0.00 (mean ± s.e.) (Fig. 4.3a). This pattern was also observed for the two <1000 mm TL P. microdon monitored in Snag Pool, which were omitted from analyses due to too few of transmissions. Tag 13 showed a depth distribution more similar to P. microdon of 1400 to 1650 mm TL (Fig. 4.3a).
Depth profiles of estuarine P. microdon 1400 to 1650 mm TL were far less skewed, and occupied depths between 0.2 ± 0.10 (mean ± s.e.) and 1.6 m ± 0.10 (mean ± s.e.) for 90 to 95 % of their time (Fig. 4.3b). The depth profile of tag 9, a 1576 mm TL P. microdon monitored in Langi Crossing during the wet season (March 2008), was similar in shape to P. microdon of the same size class, but had a greater range and maximum depth. Tag 9 spent 90 to 95 % of its time at depths between 1.2 and 4.2 m.
Pristis microdon >1850 mm TL displayed bimodal depth profiles (Fig. 4.3c). An
1883 mm TL P. microdon recorded in Milli Milli had modes at 0.7 and 2.4 m, and occupied depths of between 0.2 and 2.6 m for 90 to 95 % of the time that it was monitored (Fig. 4.3c). In the freshwater pools, two P. microdon >1850 mm TL had modes at 1.2 and 2.2 m (tag 17, Camballin Pool) and 0.4 and 2.4 m (tag 16, Myroodah
Crossing) (Fig. 4.3c). Pristis microdon with tag 17 and 16 spent 90 to 95 % of their time at depths between 0.7 and 2.9 m and 0 and 2.9 m, respectively, in the dry season
(Fig. 4.3c). Depths occupied by these two P. microdon were significantly different (p
< 0.001) between the dry and early wet season (i.e. December 2008). Maximum depths occupied by tags 17 and 16 increased from 4 and 3.1 m in the late dry season, to 6.4 and 5.3 m in the early wet season, respectively (Fig. 4.4a, e). Maximum depths increased temporarily during the peak of a ‘flood event’ (for the purposes of this study
121 a flood event is defined as: a sudden substantial increase in river discharge, which results in a >0.5-m rise in stage height).
Depth occupied by P. microdon also changed between hours of the day.
Pristis microdon was found to typically move into shallow water at sunset and from shallow water at sunrise. Times of ascent did vary between P. microdon in the freshwater pools and the estuarine pools. In general, P. microdon in the estuarine pools were observed to inhabit significantly greater depths between 06:00 and 11:00 h, and shallower depths between 15:00 and 04:00 h (p < 0.001, d.f. = 23) (Fig. 4.5a- c). Pristis microdon in the freshwater pools, inhabited significantly greater depths between 09:00 and 17:00 h, and shallower water between 18:00 and 05:00 h (p <
0.001, d.f. = 23) (Fig. 4.5a-c). The diel pattern was less exaggerated by P. microdon
<1000 mm TL. Pristis microdon of this size class generally moved into shallower water earlier (12:00 to 14:00) than larger conspecifics (Fig. 4.5a).
There was a strong positive correlation between mean hourly depth and light intensity (r = 0.86 and 0.9, p < 0.001, n = 24), but not with water temperature (r =
0.32 and 0.28, p > 0.1, n = 24) in the freshwater pools. Hourly depths of individuals displaying crepuscular movements in the estuarine pools were on average moderately and positively correlated with light intensity (r = 0.675 ± 0.082 (mean ± s.e.), p <
0.05, n = 24). Only tag 18 (1883 mm TL) in Milli Milli varied from this (p = 0.114, n
= 24). Of these individuals, mean hourly depth was moderately and negatively correlated with water temperature (r = -0.627 ± 0.096 [mean ± s.e.], p < 0.05, n = 24), with exception of tag 2 (1611 mm TL) in Snag Pool (p = 0.523, n = 24).
122 45 < 1000 mm TL 40 5 (Snag P.) 35 7 (Snag P.) 30 14 (Tele. P.) 25 13 (Tele P.) 20 15 10
Percentage of total hits total of Percentage 5 0 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 18 1400 - 1650 mm TL 15 1 (Milli) 2 (Snag P.) 12 3 (Snag P.) 11 (Tele. P.) 9
6
3
Percentage of total hits total of Percentage 0 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 18 16 >1850 mm TL 14 16 (Myr. Cr.) 17 (Camb. P.) 12 18 (Milli) 10 8 6 4 2
Percentage of total hits total of Percentage 0 0 1 2 3 4 5 6 Depth (m)
Fig. 4.3. Depth profiles of Pristis microdon (a) <1000 mm, (b) 1400 to 1650 mm and (c) >1850 mm
TL in the dry season. Only tags with >30 hits per hour for every hour are displayed. Note that the scales of the x and y-axes vary between graphs.
123 0 18 0 18
2 16 2 16
4 14 4 14
Tag Depth (m) DepthTag 6 12 6 12
Stage height (m) height Stage
8 10 8 10 Nov-2008 Dec-2008 Jan-2009 Feb-2009 Nov-2008 Dec-2008 Jan-2009 Feb-2009
0 18 0 18
2 16 2 16
4 14 4 14
Tag depth (m) depth Tag 6 12 6 12
Stage height (m) height Stage
8 10 8 10 Nov/2008 Dec/2008 Jan/2009 Feb/2009 Feb-2008 Mar-2008 Apr-2008 May-2008
0 18
2 16
4 14
Tag depth (m) depth Tag 6 12
Stage height (m) height Stage
8 10 Nov/2008 Dec/2008 Jan/2009 Feb/2009 Mar/2009 Apr/2009 May/2009 Jun/2009 Date
Fig. 4.4. Recorded depths (dots) of Pristis microdon tag (a) 17 (Camballin Pool and Myroodah
Crossing), (b) 18 (estuarine pools), (c) 19 (estuarine pools), (d) 9 (estuarine pools) and (e) 16
(Myroodah Crossing) in comparison to river stage height in the wet season. No other tags were detected during the wet season.
124 a) 0.0
0.5
1.0
1.5
Depth(m)
2.0 5 (Snag Pool) 7 (Snag Pool) 13 (Telegraph Pool) < 1000 mm TL 14 (Telegraph Pool) 2.5 b) 0 2 4 6 8 10 12 14 16 18 20 22 0.0
0.5
1.0
1.5
Depth(m)
2.0 1 (Snag Pool) 2 (Snag Pool) 3 (Snag Pool) 1400 to 1650 mm TL 11 (Telegraph Pool) 2.5
c) 0 2 4 6 8 10 12 14 16 18 20 22 0.0 16 (Myroodah Crossing) 17 (Camballin Pool) 0.5 18 (Milli Milli)
1.0
Depth (m) Depth 1.5
2.0
> 1850 mm TL 2.5
0 2 4 6 8 10 12 14 16 18 20 22
Hour
Fig. 4.5. Mean hourly depths (± s.e.) of (a) <1000 mm, (b) 1400 to 1650 mm and (c) >1850 mm TL
Pristis microdon in the Fitzroy River.
125 4.3.4 Inter-pool movement
Inter-pool movement of P. microdon was scale dependent and varied between size, region (i.e. estuarine and freshwater pools) and season. In the estuarine pools during the dry season, P. microdon of between 1400 and 1650 mm TL were only observed to move between pools in the presence of tidal waters. Conversely, P. microdon <1000 mm TL readily moved between Milli Milli and Snag Pool through a shallow
(maximum depth = 0.5 m) 0.7-km run when tidal waters were absent. However, the directions of inter-pool movements of P. microdon <1000 mm TL, were almost exclusively (98 %) the same as the water flow when in the presence of tidal waters, unlike their larger conspecifics. Only one P. microdon of this smaller size class was observed to move downstream during an incoming tide. Pristis microdon of between
1400 and 1650 mm TL were observed to move with (50 % of recorded inter-pool movements) and against (50 % of recorded inter-pool movements) the flow of water.
During the dry season in the estuarine pools, short-term (i.e. hours to days) movements of P. microdon <1650 mm TL appeared to not be in a specific direction, with individuals moving short distances up and downstream (Fig. 4.2). However, long-term (i.e. months) movements of these individuals appeared to be mostly in an upstream direction during the dry season (Fig. 4.2). Tagged P. microdon of this size class were observed to move between Milli Milli and Snag Pool in June through
August, between Snag Pool and Telegraph Pool in June through September and into
Langi Crossing in late September and November (Fig. 4.2).
Long-term inter-pool movements of P. microdon >1850 mm TL, which were tagged in the estuarine pools, was less defined. However, P. microdon >1850 mm TL tagged in Snag Pool in October and November 2008, were recorded to occasionally inhabit the downstream Milli Milli. Final transmissions for these individuals occurred
126 at Milli Milli (the only active estuarine receiver at that time) in late November and
December 2008 (Fig. 4.2). Only tag 18 was redetected at Milli Milli the following spring.
In the freshwater pools, no inter-pool movement was observed in the dry season. Two P. microdon >1850 mm TL tagged in the freshwater pools were recorded to have moved far downstream in the wet season (Fig. 4.2). Deployed in
Myroodah Crossing, tag 16 was observed to have moved 118 km downstream to Milli
Milli between 20 January and 9 April 2009, and was last detected at Milli Milli on 13
May 2009 (Fig. 4.2). Tag 17, which was deployed in Camballin Pool, moved 33 km downstream to Myroodah Crossing between 27 December 2008 and 12 January 2009
(Fig. 4.2). No further transmissions were recorded for this tag.
Detections of all P. microdon monitored in the 2008 or 2009 wet season (tags
9 and 16-19) ceased, at least temporarily, during the onset of a flood event (Fig. 4.4a- e). Only tag 16 was detected again in the same pool it inhabited prior to the flood event, and only temporarily when stage height decreased to near dry season levels.
Tags 16 and 17 were detected to have moved into another pool after a flood event.
4.3.5 Relative abundance of Pristis microdon prey
Seine netting conducted on shallow sandbars in the freshwater pools in the early and late dry seasons in 2007 and 2008, produced 24 fish species and two species of crustaceans: M. rosenbergii and an unidentified freshwater shrimp. Only six species were present in the majority (>50 %) of seine samplings. These included the barred grunter (Amniataba percoides), Prince Regent hardyhead (Craterocephalus lentiginosus) and flathead goby (Glossogobius giuris), as well as known P. microdon prey, including M. rosenbergii, N. erebi and N. graeffei (Thorburn et al. 2007;
127 Peverell 2008). Thirty-four percent of the total number of captured fishes were >50 mm TL. Each of the six most abundant species measured, with the exception of C. lentiginosus, were observed to reach a size of >50 mm TL. All known P. microdon prey species that were captured were observed to reach a size of >100 mm TL.
Glossogobius giuris, M. rosenbergii, N. erebi and N. graeffei showed a significant increase in relative abundance in night seines (p < 0.05, n = 5 to 8). Ten species: M. rosenbergii, toothless catfish (Anodontiglanis dahli), mouth almighty (Glossamia aprion), diamond mullet (Liza alata), barramundi (Lates calcarifer), western rainbowfish (Melanotaenia australis), ox-eye herring (Megalops cyprinoides), black catfish (Neosilurus ater) and Hyrtil’s tandan (Neosilurus hyrtlii) were only captured in night seines. Only four species: Prince Regent hardyhead (Craterocephalus lentiginosus), western sooty grunter (Hephaestus jenkinsi), spotted butterfish
(Scatophagus argus) and freshwater longtom (Strongylura krefftii) were only captured in day seines.
4.4 Discussion
This study is the first to document the vertical and inter-pool movements of juvenile
P. microdon within a riverine nursery environment. Results from this study demonstrated P. microdon habitat use patterns to vary temporally, spatially and ontogenetically. As hypothesised, patterns of habitat use appeared to vary in response to changes in environmental variables within the system. Significant findings of this study, including those concerning the utility of passive acoustic monitoring within the
Fitzroy River, are expanded upon below. It should be noted that the results described
128 below are based off a small sample size and not all behaviours may be representative of the population as a whole.
4.4.1 Depth use
Pristis microdon in the Fitzroy River displayed ontogenetic depth partitioning, with larger individuals occupying deeper water. Differences in depth use of fishes have been attributed to variations in trophic (Ebert 2002; Collins et al. 2005; McElroy et al.
2006; Grubbs 2010) and habitat requirements (e.g. predator avoidance) (Grubbs 2010;
Simpfendorfer et al. 2010). Young of the year P. microdon are likely to face a greater risk of predation than larger size classes and require habitat that aid in reducing this threat. Occupying shallow waters would likely provide YOY with sanctuary from larger bodied predators, like the bull shark (Carcharhinus leucas). In addition, the warmer shallow waters would also likely increase the metabolism, and resultantly growth, of YOY (Morrisey and Gruber 1993; Heupel et al. 2007; Grubbs 2010;
Simpfendorfer et al. 2010). Obtaining a larger size faster would aid in decreasing the risk of predation on YOY. In contrast, larger bodied P. microdon are less manoeuvrable in shallow water and may increase their ability to escape predators and/or hunt prey by moving into deeper water. The observed ontogenetic depth partitioning would reasonably be advantageous to all size classes as it would likely decrease intra-specific competition between these groups.
Depths occupied by larger P. microdon appeared not to be associated with a particular pool depth but rather a specific environment. In the assumption that a particular depth was preferred by P. microdon, the depths occupied in the wet season would be similar to those occupied in the dry season. Similarly, the depths occupied by a size class would not vary between pools. However, P. microdon was found to
129 inhabit significantly greater depths in the wet season as stage height increased, and when occupying deeper pools. The observed increases in depth of P. microdon may be due to movement of individuals to deeper water. Alternatively, or additionally, the increase in depth may have been a direct result of the increase in stage height, rather than movement of P. microdon to another area. Depths occupied by P. microdon during the day in the dry and wet seasons were fairly consistent with bottom depths in the freshwater pools (see Chapter 2). This may suggest that larger P. microdon have an affinity for the riverbed and that the increase in depth was, at least in part, a result of an increase in stage height. The relatively level riverbed has abundant coverage in the form of snags and overhangs and is likely to provide a suitable resting space for these benthic rays.
Depth of most size classes was not homogenous through the day and varied temporally at an hourly scale. In general, P. microdon occupied shallower waters at night and deeper water in the day. Movement into and out of shallow water transpired around sunset and sunrise, respectively. Only YOY P. microdon varied from this pattern and moved into the extreme shallows in the afternoon. Reasoning for this different pattern is unclear but would expose YOY to maximum daily temperatures
(Chapter 2), and thus this behaviour may result in increased growth of YOY
(Morrisey and Gruber 1993; Grubbs 2010). Crepuscular movements similar to those observed in this study are not uncommon and have been documented in other elasmobranchs, including the mostly benthic small-spotted catshark (Scyliorhinus canicula) (Sims 2003), school shark (Galeorhinus galeus) (West and Stevens 2001) and Pacific sleeper shark (Somniosus pacificus) (Hulbert et al. 2006). Often these diel movements develop from the need of fish to move between habitats to access resources, such as prey and cover (Holland et al. 1993; Cartamill et al. 2003), or as
130 means of behavioural thermoregulation (Matern et al. 2000; Sims et al. 2006; Hight and Lowe 2007). Movement of P. microdon into shallow waters at night, places these individuals in the space and time when and where relative abundance of known (i.e.
M. rosenbergii, N. erebi and N. graeffei) and potential (i.e. abundant species of reasonable size) prey species increases.
Diel vertical and horizontal movements are known to be triggered by cues, such as light, temperature, physiological state (e.g. hunger) and/or endogenous rhythmicity (i.e. internal clock) (Nelson and Johnson 1970; Gibson 1997; Gruber et al. 1998). Crepuscular movement into and out of shallow depths suggests light, or the strongly correlated temperature, to be the primary cue for the vertical movements of juvenile P. microdon (West and Stevens 2001; Cartamill et al. 2003; Hulbert et al.
2006; Dewar et al. 2008). Discerning whether crepuscular movement is based off of light or temperature can be difficult due to their close relationship, as was the case in the estuarine pools. However, only light intensity was found to be at least moderately correlated with hourly depth of P. microdon in the estuarine and freshwater pools. It is thus reasonable to hypothesise that light is the primary cue for P. microdon vertical movements. However, further research is needed to verify this hypothesis and to explain the difference between the timing of P. microdon vertical movements in the estuarine and freshwater pools.
4.4.2 Inter-pool movement
Inter-pool movements of P. microdon appeared to be influenced by water depth and flow and varied between size classes and seasons. In the estuarine pools during the dry season, short-term (i.e. hourly to daily) movements between pools were equally in an upstream and downstream direction. Movement between pools appeared less
131 restricted for YOY, which were observed to move unrestrained through shallow runs connecting pools during low tide, unlike their larger counterparts. Other riverine species have been shown to take advantage of their ability to access areas unavailable to larger competitors and predators to escape situations of high biotic pressure (Power et al. 1985; Schaefer 2001). Similarly, this ability may allow YOY to move into areas where there is less competition from larger P. microdon or areas with lower risk of predation. However, the direction of YOY inter-pool movements was restricted, or at least influenced by the direction of tidal flow, unlike with larger P. microdon. This could suggest that YOY are not physically capable of moving against such high flows.
Alternatively, this could suggest that movement with the current is beneficial for
YOY, as observed with the leopard shark (Triakis semifasciata) (Ackerman et al.
2000; Carlisle and Starr 2009), sandbar shark (Carcharhinus plumbeus) (Medved and
Marshall 1983), dwarf sawfish (Pristis clavata) (Stevens et al. 2008) and cownose ray
(Rhinoptera bonasus) (Smith and Merriner 1985). Movement with tidal flow has been suggested to be a means of decreasing energetic costs (Ackerman et al. 2000) and maximizing foraging efforts (Smith and Merriner 1985; Ackerman et al. 2000).
The argument that YOY are physically capable of swimming against tidal flow, at least for short distances, was supported by the observation of a YOY that moved downstream during an incoming tide.
Long-term (seasonal) movements of P. microdon were more directional and dependent upon season as well as size. In the dry season, pooled results of P. microdon <1650 mm TL in the estuarine pools demonstrated P. microdon of this size class to move slowly in an upstream direction, away from increasing salinity and daily temperature fluctuations. Euryhaline C. leucas have also been noted to move in accordance with changes in salinity, moving towards the river mouth during high
132 inflows of freshwater (Heupel and Simpfendorfer 2008). Heupel and Simpfendorfer
(2008) concluded that such behaviour may be a means of reducing the costs of osmoregulation. As a euryhaline species, juvenile P. microdon may be able to tolerate marine water, but as a freshwater resident, occupying high saline water may be too taxing and energetically costly for juvenile P. microdon <1650 mm TL.
Movement upstream into the relatively environmentally stable freshwater pools
(Chapter 3) may be a form of behavioral regulation/homeostasis. Although a large percentage of YOY were found to remain in the estuarine pools throughout the dry season in certain years, this appears to be a result of decreased access to upstream pools. Only in years with relatively large wet seasons was a large portion of YOY observed to move upstream (Chapter 3). In years with small, short and/or early wet seasons, a larger percentage of YOY would likely be exposed to lower water levels, cooler temperatures and decreased flows (i.e. early dry season conditions) sooner after parturition. Decreased water levels would decrease the accessibility of YOY into upstream pools, leaving many trapped in the lower reaches of the river. Additionally, cooler water temperatures, which have been demonstrated to reduce movement in P. microdon (Peverell pers. comm.), and potentially decreased flows, which have been shown to decrease upstream movement of other riverine fishes (Banks 1969), may reduce any ‘instinct’ for YOY to move upstream.
Long-term wet season movements were only observed for P. microdon >1850 mm TL, but demonstrated that a number of P. microdon of this size class move downstream into estuarine waters during the wet season. This was most evident with the observation of a P. microdon that moved 118 km downstream into the river mouth area. Peverell (2008) also noted similar movement of P. microdon during the wet season after recapturing a 1940 mm TL P. microdon in an estuary after it had been
133 tagged 100 km upstream. Movement of P. microdon into estuarine and/or marine waters at sizes of >1850 mm TL was supported by captures of P. microdon of between 2100 and 2600 mm TL in tidal creeks, estuaries and nearshore waters of
King Sound (Thorburn et al. 2007; Chapter 3). Similarly, Simpfendorfer et al. (2010) noted smalltooth sawfish (Pristis pectinata) >1800 mm TL to move from shallow mangrove-lined areas to habitats in the marine fringe of an estuary, and eventually into deeper waters.
4.4.3 Acoustic tag performance and utility
Acoustic transmitters and receivers were selected for this study as they have been shown to be effective in tracking animals within linear flow environments, including rivers, river mouth areas/estuaries and sloughs (Caron et al. 2002; Tavemy et al.
2002; Clements et al. 2005; Heupel and Simpfendorfer 2008; Simpfendorfer et al.
2008a; Carlisle and Starr 2009; Jellyman 2009). Similarly, passive acoustic monitoring proved to be a valuable method in monitoring the movements of P. microdon during the dry season, in the estuarine and freshwater pools of the Fitzroy
River. Detection range/power of the tags deployed on P. microdon also appeared adequate for the implemented array design. Transmission failure of tags in shallow waters can be an issue in tracking shallow water species. However, the abundance of transmissions from depths of <0.5 m demonstrated that VR2W receivers were capable of detecting the deployed acoustic tags in extreme shallow waters. Although, the distance to which the acoustic signals travel through such shallow water was unclear.
Detection of transmissions was less reliable in the wet season. Transmissions were only periodically detected in times of relatively low flow during the wet season.
A reduction in the number of recorded transmissions in high flows was expected, as
134 high suspended sediment loads, ambient noise and water entrained bubbles are known to degrade acoustic transmissions (Klimley et al. 1998; Voegeli et al. 1998; Lacroix et al. 2005; Heupel et al. 2006). Degradation of acoustic signals may have resulted in the cessation of detections observed at the onset of flood events, giving the false appearance of individuals moving out of their respective pools. However, this was proven to not be the case for at least two P. microdon, which were detected in pools downstream of their release sites after the cessation of transmissions occurred at their tagging site.
External attachment of acoustic tags, via the use of Rototags, proved to be a quick and useful method for the deployment of tags that were intended to last for several months. Duration of the monitored periods of YOY P. microdon tagged using this method were typically shorter than for larger P. microdon. The shorter monitoring period of YOY may be due to an increased rate of tag shedding (a result of a faster rate of growth) (Stevens et al. 2000; Chapter 3), a higher rate of mortality
(Chapter 3), which is common with YOY elasmobranchs (Cortés and Parsons 1996;
Simpfendorfer 2000; Heupel and Simpfendorfer 2002) and/or the ability of YOY to move more readily out of monitored pools than larger individuals. Although this method of attachment may be useful for the monitoring of larger individuals, internal implantation, a proven method used in the monitoring of elasmobranchs (McKibben and Nelson 1986; Holland et al. 1999; Sims et al. 2001), may prove more effective in documenting movements of at least YOY. Preliminary results from a riverine acoustic telemetry study, which implanted acoustic tags in P. microdon, has shown promising results and observed complete healing of subjects within several weeks
(Peverell pers. comm.).
135 No negative health issues were observed to develop as a result of tagging methods used in this study. This is supported from the daily movement of all tagged
P. microdon through to their tags’ final detection (i.e. no P. microdon were observed to have died in monitored pools). Also, recapture of three tagged P. microdon on hook and line, one day, several months and three years after being tagged (Chapter 5), suggests that these animals were healthy enough to feed at those times.
4.4.4 Conclusion
Results of this study showed that depth use and/or inter-pool movements of juvenile
P. microdon vary spatially, temporally and ontogenetically. These changes appeared to be influenced by variations in river depth, flow and light intensity, although these effects varied between size classes. Prey distribution, salinity and water temperature are also believed to influence these movements, but further research is required to confirm these hypotheses. As such, changes to these environmental variables brought on by anthropogenic or climatic factors are likely to also influence the movement and habitat use of juvenile P. microdon. Anthropogenic influences may have a negligible effect on the habitat use patterns of P. microdon during average years. However, caution needs to be used during years of extreme environmental fluctuations, as anthropogenic influences may compound factors that occur naturally. In years of low precipitation for example, restriction of river flow and or the abstraction of water from the river would further reduce water levels and thus inter-pool movement of P. microdon.
Future research is needed to aid in furthering the understanding of environmental influences on P. microdon habitat use. Specifically, an increase in sample size is needed to increase the power of statistical analyses. In addition,
136 tagging of neonates through the wet season prior to their movement upstream would aid in exploring how changes in flow, stage height and temperature influence upstream movement of YOY. Lastly, it is the recommendation of the author that future research investigate the efficacy of acoustic methods under extreme wet season conditions.
137 Chapter 5: Utility of conventional and satellite tags in
the monitoring of large-scale movement and habitat
use patterns of the freshwater sawfish Pristis
microdon and northern river shark Glyphis garricki
5.1 Introduction
Knowledge of large-scale movements of elasmobranchs can be valuable in the understanding of a species habitat use patterns throughout its home range and life stages (Stevens et al. 2000; Eckert and Stewart 2001; Heithaus et al. 2007b; Carlson et al. 2008). The euryhaline freshwater sawfish (Pristis microdon) and northern river shark (Glyphis garricki) have been observed to occupy a diversity of riverine, estuarine and marine environments (Taniuchi et al. 1991; Thorburn and Morgan 2004;
Peverell 2005; Thorburn et al. 2007; Pillans et al. 2009; Chapter 3). However, little is known about how these species use such areas. This can be attributed to the recent discovery of G. garricki in Australia (Taniuchi et al. 1991; Pillans et al. 2009) and the apparent rarity of P. microdon and G. garricki. Additionally, the monitoring of these species can be a daunting task, due to the remoteness and extremity of the environments in which they inhabit. To obtain the data necessary for the development of effective conservation strategies, appropriate methods for large-scale tracking of these critically endangered species need to be established.
Mark and recapture (e.g. Stevens et al. 2000; Kohler and Turner 2001; Kohler et al. 2002; Queiroz et al. 2005) as well as electronic tagging methods (e.g. Arnold and Dewar 2001; Eckert and Stewart 2001; Weng et al. 2008) have been shown to be
138 of great use in the monitoring of large-scale movement patterns of elasmobranchs.
Satellite telemetry has proven to be of high value in monitoring the movements of wide ranging species, like the great white shark (Carcharadon carcharius) (Bruce et al. 2006; Weng et al. 2007), basking shark (Cetorhinus maximus) (Priede 1984; Sims et al. 2003; Southall et al. 2005), blue shark (Prionace glauca) (Stevens et al. 2010) and salmon shark (Lamna ditropis) (Weng et al. 2005, 2008), as it is not restricted to the detection range of stationary receivers (see Chapter 4). Instead, many satellite tags rely on low orbiting Argos satellites to detect radio transmissions emitted by tags.
However, the aerials of the tags must be out of the water to successfully transmit, as radio transmissions produced by satellite tags do not have the power to penetrate far through the water column. Conductivity switches are thus used in satellite tags to help the tag differentiate between surface and subsurface environments, and signal when to transmit. Unfortunately the application of tags with conductivity switches is limited to marine environments as the tag would be misled to continuously transmit in freshwater.
Mark and recapture methods, on the other hand, are not restricted by battery life, success of transmissions or to areas monitored by receivers. However, this method is dependent upon recapture of tagged individuals and thus limited to the location and intensity of fishing effort. Conventional tags involved in mark and recapture studies are inexpensive and can be deployed on a large sample of the population, unlike most electronic tags.
It was the aim of this study to deploy conventional and satellite tags on P. microdon and G. garricki to document the large-scale movements of these species in the Fitzroy River and King Sound, Western Australia, to better understand the habitat use of these species. Additionally, this study aimed to assess the utility of
139 conventional and satellite tags in the monitoring of P. microdon and G. garricki throughout their home range. It was hypothesised that satellite telemetry would be an effective means to monitor the large-scale movements of P. microdon and G. garricki, as these species are known to inhabit shallow water and P. microdon has been observed to expose its dorsal fin when foraging (Thorburn and Morgan 2004;
Thorburn 2006; Thorburn et al. 2007; Pillans et al. 2009; Chapter 4). In addition, it was hypothesised that mark and recapture methods would atone for the limitations of satellite and acoustic telemetry in documenting large-scale movements of the studied species (see Chapters 1, 4).
5.2 Material & Methods
5.2.1 Study Site
Sampling, tagging and tracking efforts of P. microdon and G. garricki were undertaken in the lower 160 km of the Fitzroy River and within King Sound, Western
Australia in 2005 to 2009 (see Fig. 5.1). The study area experiences monsoonal wet and dry seasons as well as large tides, which cause abiotic variables such as temperature, salinity and turbidity to vary significantly throughout the year (see
Chapter 2). During the wet season, which typically spans between December and
April, the study area experiences a substantial increase in precipitation. At this time, river discharge and stage height significantly increase (see Chapter 2), allowing pools previously isolated during the dry season to reconnect. During the dry season, the river is obstructed by road crossings and the Camballin Barrage, a small weir.
However, floods can be large enough to allow fish movement over these obstacles in the wet season. Although most of the river is fresh (<1.0 ppt), the lower estuarine
140 pools (0 to 17.5 km from the mouth) are transformed from a freshwater to a marine environment during the dry season (Chapter 2).
King Sound is primarily fed by the Indian Ocean and is bordered by numerous tidal creeks and five relatively large rivers (i.e. the Fitzroy, Fraser, May, Meda and
Robinson River (see Fig. 5.1). King Sound experiences mixed macro-tides (11+ m), which produce large currents as well as highly turbid and well mixed waters in the southern nearshore areas (see Chapter 2). During extreme low tides the upper half of the sound is drained, leaving only narrow channels filled with water. Low tides also often drain the tidal creeks bordering the upper King Sound, exposing mud flats and forming isolated pools within the creeks.
141 32
Fig. 5.1. Survey sites in the Fitzroy River and King Sound, Western Australia.
5.2.2 Tagging
Pristis microdon were captured in the Fitzroy River between 2003 and 2009
(including data from previous studies; see “Data Analysis”), with 20-m gillnets consisting of 51, 76, 102, 152 and 203-mm stretched monofilament and by hook and line. Glyphis garricki (n = 7) and a single P. microdon were also captured in King
Sound between 2007 and 2009, with 20 and 50-m gillnets consisting of 152 and 178- mm stretched monofilament, respectively. In the river, the majority of P. microdon were moved to the shore-line to be measured and tagged, ensuring that their spiracles,
142 gills and mouth remained submerged to allow for respiration. Pristis microdon <1200 mm total length (TL) were occasionally measured and tagged aboard the research vessel to reduce overall handling time. Elasmobranchs <2000 mm TL that were captured in King Sound were measured and tagged aboard a research vessel. A 25- mm diameter hose was used to pump ambient water through the mouth and gills of fishes measured onboard the vessel. Larger animals captured in King Sound were not brought aboard to prevent injury to the animal and to researchers. Upon capture, specimens were inverted on to their dorsal surface, which caused them to enter tonic- immobility. Pristis microdon and G. garricki were then measured, sexed and tagged.
Conventional tagging occurred in both the Fitzroy River and King Sound.
Conventional tags were deployed in estuarine pools of the river, which consisted of
Snag Pool (6.2 km upstream of the river mouth), Telegraph Pool (11.6 km upstream of the river mouth) and Langi Crossing (14.9 km upstream of the river mouth), as well as in freshwater pools, which consisted of Myroodah Crossing (123 km upstream of the river mouth), Camballin Pool (156.5 km upstream of the river mouth; name assigned by the author) and Lower Barrage Pool (160.7 km upstream of the river mouth; name assigned by the author) (Fig. 5.1). Tags were only deployed on healthy specimens. Individually numbered Rototags (Dalton Supplies, Woolgoolga, New
South Wales, Australia) were attached to the anterior section of either the first or second dorsal fin, following procedures similar to those described by Heupel et al.
(1998) (Fig. 5.2 a). Tags were labeled with the contact information of the research organisation to provide fishers with a means to report recaptured fishes. One hundred twenty-four P. microdon ranging in size between 763 to 2770 mm TL were tagged between 2003 and 2009 by this and previous studies (Thorburn and Morgan 2004;
Thorburn 2006; Thorburn et al. 2007; Chapter 3). Only three G. garricki were tagged
143 with Rototags between 2007 and 2009. No G. garricki recaptures were reported or observed and therefore are not reported on further.
Smart Position or Temperature Transmitting Tags (SPOT-5) (Wildlife
Computers, Redmond, Washington, USA) set in fin-mount molds, were deployed on
P. microdon and G. garricki (see Fig. 5.2 b) captured within King Sound and Snag
Pool, Fitzroy River in June, July and October 2007 (Table 5.1). These tags were set to only transmit at 13 selected hours throughout the day to maximise battery life.
Transmission times of tags deployed in June and July were intended to maximise the chance of transmission (i.e. set to transmit when fishes were hypothesised to be most active in shallow water). For P. microdon, nine of the 13 transmission times were between 17:00 and 04:00 h. As G. garricki behaviour was unknown, five daytime
(08:00 to 16:59 h) and nighttime (20:00 to 04:59 h) periods and three periods surrounding sunrise and sunset (05:00 to 07:59 h and 17:00 to 19:59 h) were used for tag 41103, deployed in June 2007. Transmission times of tags deployed in October
2007 (tags 41101 and 41099), were set to maximise the chance of satellite reception.
This was based upon when Argos satellites were most commonly overhead, as determined using SatScape V2.02 (www.satscape.co.uk). Tagged individuals were selected by the size of the animal to ensure the tag weight was <1 % of the animal’s weight. The SPOT-5 tags were attached to the first dorsal fin of each specimen using four nylon bolts (4 mm in diameter), which were secured using nylon lock washers and stainless steel nuts and washers. Four holes, which securing bolts were passed through, were created with implant-grade stainless steel drill bits (5 mm in diameter).
Proper placement of the holes was established with the use of a prefabricated drilling template.
144 Location estimates of tags were calculated using the transmissions received by
Argos satellites in association with known statistics (e.g. time and distance from last transmission, maximum swimming speed of tagged subject). Location estimates were downloaded from the Argos website (www.argos-system.org). Accuracy of location estimates was dependent upon the number of transmissions received during a single satellite pass. A greater number of transmissions produced a greater accuracy estimate score. Locations were graded upon their estimated accuracy, with scores of
0, 1, 2 and 3 having a >1500, <1500, <500 and <250-m radius, respectively.
Accuracy estimates for scores of A or B were not given due to too few of transmissions, but may have been highly accurate. Recorded transmissions were checked for accuracy through examination of the area covered by the satellite at time of tag transmission, as single erroneous transmissions can be caused by noise or corrupted messages from other satellite tags. If the satellite footprint (area monitored by the satellite at a given time) covered the King Sound region or nearby areas, it was likely to have come from tags deployed by this study and was accepted.
Table 5.1. Satellite tag deployment information
Species TL (mm) Sex SPOT tag # Release date Capture location P. microdon 1780 F 41104 03-Jul-07 Snag Pool P. microdon 2580 M 41099 24-Oct-07 Point Torment G. garricki 1365 M 41103 19-Jun-07 Doctor's Creek G. garricki 1084 F 41101 18-Oct-07 Robinson River estuary
145
Fig. 5.2. (a) Rototag and (b) SPOT-5 satellite tag attachment sites and methods.
5.2.3 Data analysis
Maps were produced using ArcGIS 9.3.1 (ESRI, Redlands, California, USA). Data analysed in this study was collected between 2005 and 2010 (2010 data consisted only of recapture information reported by fishers). However, tagging data acquired in
2003 and 2004 by Thorburn et al. (2007) was included in analyses as P. microdon tagged during this period were recaptured during the study period (i.e. 2005 to 2010).
Sampling during this previous research was conducted during similar months and at similar sites (within the lower 160 km of the river) as this study. Size classes of P. microdon were divided by their approximate ages following Peverell (2008) and
Thorburn et al. (2007).
In analysing the data, results from P. microdon recaptures and acoustic telemetry data (see Chapter 4) were treated independently, even though acoustic tags and Rototags were attached to the same individuals. In this study 22 conventionally- tagged P. microdon were also fitted with acoustic tags. However, only five of these individuals were recaptured, with three being recaptured after being at liberty for <1 day (Table 5.2). Thus there was only minor overlap between data from the two methods.
146 Surface times of SPOT-5 tagged fishes were estimated from times in which transmissions from satellite tags were detected. The number of transmissions for day hours (08:00 to 16:00 h), night hours (20:00 to 04:00 h) and hours surrounding sunrise and sunset (05:00 to 07:00 h and 17:00 to 19:00 h), as well as for incoming and outgoing tides were compared to explore if depth use (i.e. surface time) varied between these periods. However, bias was introduced into these results when transmissions times of tags were preset, due to variability in the number of satellite passes that occur at different hours and the different tidal states (the duration of incoming tides are nearly two hours shorter than that of outgoing tides). To decrease this bias, the number of transmissions for each time period and tidal status was divided by the number of satellite passes to occur during these respective times.
5.3 Results
5.3.1 Rototags
Pristis microdon
Thirty-three of the 124 P. microdon tagged between 2003 and 2009 (excluding one recaptured P. microdon that had shed its tags, as was evident by a fresh tagging scar), were recaptured and/or resighted by researchers and/or fishers between one and four times, totaling 49 recaptures, between 2005 and 2010 (Table 5.2). Time between tagging and final recapture ranged between <1 to 1246 days (Table 5.2). Only 14 individuals were recaptured after being at liberty for five or more days between captures (Table 5.2). Of these, tag 7092 was recaptured twice after being at liberty for greater than five days after each release (discussed below). As such, there was a total of 15 relatively ‘long-term’ recaptures in this study. The 14 P. microdon recaptured
147 after more than five days consisted of six (out of 64 tagged) <1150 mm TL (YOY), two (out of seven tagged) 1320 to 1611 mm TL (1+ to 2+ year old) and six (out of 53 tagged) 1780 to 2770 mm TL (2+ to 5+ year old) P. microdon. Final recaptures were also recorded in 2003 and 2004 by Thorburn et al. (2007) for four additional P. microdon (2150, 2200, 2240 and 2262 mm TL) that were at liberty for 140, 1, 116 and
116 days, respectively, in freshwater pools during the dry season. Additionally,
Thorburn et al. (2007) reported tag F7056, also captured in 2005 (Table 5.2), to be recaptured twice in 2004 (96 and 20 days after previous release).
All P. microdon captured between 2005 and 2010, were recaptured at the site in which they were tagged, with exception of three individuals (Table 5.2, Fig. 5.3).
One female P. microdon (M1032; 977 mm TL at initial capture and 1050 mm TL at recapture) at liberty for 143 days (June to November 2008), was recaptured 9 km upstream from where it was tagged (i.e. Snag Pool) (Table 5.2, Fig. 5.3). Another female P. microdon (F7021; 2320 mm TL at initial capture and ~2540 mm TL at recapture) at liberty for 1010 days (recaptured August 2006), was recaptured in Snag
Pool, 142 km downstream from where it was tagged (i.e. Camballin Pool) (Table 5.2,
Fig. 5.3). One male P. microdon (F7092; 1611 mm TL at initial capture and ~1981 mm TL at recapture), at liberty for 989 days (recaptured March 2010), was recaptured near the river mouth (i.e. the Cuttings), ~10 km downstream from where it was tagged
(i.e. Snag Pool) (Table 5.2, Fig. 5.3). This last individual was recaptured again in
Snag Pool, 257 days (recaptured November 2010) after its previous recapture. At this time it was reported that this fish was found with no rostrum. Of the recaptures that occurred in pools in which the animal was last released, 11 were at liberty for between five and 481 days (Table 5.2). Six of these 11 recaptures had access to other pools through increased stage heights during wet seasons (M1034, F7056), tidal assistance
148 (F9850, F9849, F8086) or general connectivity of pools at timing of recapture
(M1044).
Table 5.2. Pristis microdon recaptures in the Fitzroy River in 2005 to 2010
Tag #, Rototag number; t0, time of initial capture; tfr – t0, time at liberty; # recap., number of recaptures;
TL0, total length at initial capture; Capture method, method used for initial capture and recaptures (HL,
hook and line; CN, cast net; V, visual; #, mesh size in mm of gill net)
Capture Final recapture TL # Tag # t t - t 0 Capture method 0 location location fr 0 (mm) recap. F7092* 23-Jun-07 Snag Pl Snag Pl 1246 1611 3 HL,152,HL,HL F7021 4-Nov-03 Camballin Pl Snag Pl 1010 2320 1 51,76/152** M1034 17-Sep-09 Myroodah Crs. Myroodah Crs. 317 1865 1 102,152 F7056 15-Jul-04 Lwr Barrage Pl Lwr Barrage Pl 308 2660 1 152,HL M1032 30-Jun-08 Snag Pl Langi Crs. 143 977 1 152,HL F7081 3-Aug-06 Lwr Barrage Pl Lwr Barrage Pl 84 1150 2 152/203,152/203,152** M1045* 28-Oct-08 Camballin Pl Camballin Pl 23 2210 1 152,120 F9850 23-Jun-07 Snag Pl Snag Pl 22 1000 3 203,152,203 M1044 25-Jun-08 Myroodah Crs. Myroodah Crs. 14 879 1 152,HL F7079 13-Jul-05 Lwr Barrage Pl Lwr Barrage Pl 12 2299 2 HL,152,HL F9849 23-Jun-07 Snag Pl Snag Pl 11 1016 2 152,203,102 F8086 23-Jun-07 Snag Pl Snag Pl 10 950 3 152,152,152,152 M1093 15-Sep-09 Myroodah Crs. Myroodah Crs. 7 1580 1 CN,V M1094 17-Sep-09 Myroodah Crs. Myroodah Crs. 5 2499 2 152,203,V M1098 16-Sep-09 Myroodah Crs. Myroodah Crs. 2 1003 1 152,203 F9836 22-Jun-07 Snag Pl Snag Pl 2 1000 1 HL,203 F9838 22-Jun-07 Snag Pl Snag Pl 2 985 4 HL,102,102,152,152 F7096 29-Jun-07 Camballin Pl Camballin Pl 2 935 2 203,102,120 F9835* 22-Jun-07 Snag Pl Snag Pl 1 1555 1 152,120 M1021* 29-Jun-08 Snag Pl Snag Pl 1 1423 1 152,152 F7068 6-Aug-06 Myroodah Crs. Myroodah Crs. 1 1079 1 152,152 M1066 14-May-09 Lwr Barrage Pl Lwr Barrage Pl 1 1013 1 102,HL F9843 30-Jun-07 Camballin Pl Camballin Pl 1 900 1 102,203 F9841 23-Jun-07 Snag Pl Snag Pl 1 846 1 152,152 M1095 22-Sep-09 Myroodah Crs. Myroodah Crs. 0 1065 1 203,V F9848 23-Jun-07 Snag Pl Snag Pl 0 1024 1 152,152 M1035 1-Jul-08 Snag Pl Snag Pl 0 989 2 152,152,152 F9807 23-Jun-07 Snag Pl Snag Pl 0 941 1 152,102 F9844 4-Jul-07 Snag Pl Snag Pl 0 914 1 152,152 M1079* 14-May-09 Lwr Barrage Pl Lwr Barrage Pl 0 872 1 152,102 M1027 30-Jun-08 Snag Pl Snag Pl 0 852 1 152,152 F9809 23-Jun-07 Snag Pl Snag Pl 0 825 1 152,102 F9810 23-Jun-07 Snag Pl Snag Pl 0 822 2 152,102,102 * Fitted with an acoustic tag (see Chapter 4)
** Mesh size that sawfish were captured in was unknown, but net deployed consisted of 76 and 152 or
152 and 203 mm mesh
149 King Sound b
Western Australia
The Cuttings
17°30'S Snag Pl. ò ò
Langi Crs. 18°0'S Camballin Pl. ¸ ò 0 5 10 20 30 40
Kilometers
123°30'E 124°0'E 124°30'E ò tag F7092 ò tag F7021 ò tag M1032
Fig. 5.3. Initial capture (diamond) and recapture (circle) sites of conventionally tagged Pristis
microdon observed to move between pools in the Fitzroy River (2005 to 2010).
5.3.2 Satellite tags
Pristis microdon
Two successful transmissions from tag 41104 (deployed 3 July 2007 in Snag Pool,
Fitzroy River) were recorded to have occurred on 2 September 2007 and 12
November 2007. No location estimates were produced for this tag. Seven different
transmissions were recorded from tag 41099 (deployed 24 October 2007 at Point
Torment) on the day of release. The two location estimates calculated for this tag was
within close proximity to the deployment site (see Fig. 5.4), with 1 and B grade
accuracy. Another transmission from tag 41099 was recorded on 14 April 2008, but
150 the recording satellite’s position was above Antarctica at the time of transmission and the transmission was classified as erroneous.
Glyphis garricki
No transmissions were recorded from tag 41101 (deployed 18 October 2007, in the
Robinson River estuary). A total of 18 different transmissions from tag 41103
(deployed 19 June 2007, in Doctor’s Creek) were recorded between 22 June and 26
September 2007. These transmissions represented 11 different transmission periods
(i.e. nonconsecutive transmissions with intermissions of at least 10 min.). Seven of these transmissions occurred at night, six of which occurred between 01:00 and 03:00 h, three were during daylight hours and one occurred near sunset. When standardised
(i.e. the number of transmissions divided by the number of available satellite passes during each period of day), this came to 1.7, 0.8 and 0.5 transmissions per 100 satellite passes in night, day and sunset hours, respectively. In regards to tidal status, nine of the 11 transmissions occurred on incoming tides. When standardised (i.e. the number of transmissions divided by the number of available satellite passes during incoming and outgoing tides that occurred at preset transmission hours), this came to
1.9 and 0.4 transmissions per 100 satellite passes during incoming and outgoing tides, respectively.
For tag 41103, two location estimates, five minutes and 35 km apart, were produced on 23 June 2007 (Fig. 5.4). Accuracy estimates for these locations were given a B-grade. All location estimates of this tag were within King Sound and showed a general movement west from Doctor’s Creek (near Derby) to a creek near
Valentine Island (west coast of King Sound) (Fig. 5.1). This creek was similar in size to other areas that G. garricki were captured at in King Sound (Chapter 3). All
151 recorded transmissions from tag 41103 occurred at times in which the recording
satellite was overhead of King Sound.
16°30'S
Robinson River King Sound :
: 17°0'S
Meda River :
May River Fraser River
b 17°30'S Western : Australia ¸ Fitzroy River 0 5 10 20 30 40
Kilometers 123°0'E 123°30'E 124°0'E ò Pristis microdon ò Glyphis garricki Most direct path
Fig. 5.4. Deployment (diamond) and calculated transmission locations (circles) for tagged Pristis
microdon and Glyphis garricki in King Sound.
5.4 Discussion
Monitoring the movements of euryhaline species can be difficult due to their ability to
inhabit a wide range of environments. It was the aim of this study to employ and
152 assess mark and recapture methods along with satellite telemetry to monitor the large- scale movements of P. microdon and G. garricki in King Sound and its freshwater tributaries, including the Fitzroy River. Results from mark and recapture methods were limited but proved to be of some use for monitoring large-scale movements of P. microdon in a riverine environment. Conversely, SPOT-5 satellite tags in fin mounts unexpectedly provided too little of data to be of use for monitoring P. microdon or G. garricki movements. Although limited, results from this study were used to comment on habitat use of P. microdon and G. garricki.
5.4.1 Rototags
5.4.1.1 Utility
Movement data of P. microdon in the Fitzroy River derived from conventional tagging were expectedly limited. Only 17 P. microdon were recaptured after being at liberty for between five and 1246 days between 2003 and 2010, a 13.7 % return rate.
The 0 % return rate of tagged G. garricki was also expected, because of the limited number of tag deployments (n = 3). Low return rates are not uncommon, as illustrated by Kohler and Turner (2001) who reviewed 52 mark and recapture studies involving shark species. In this review 55 % of the reported recapture rates (n = 191), which incorporated a variety of conventional tags, including Rototags, observed return rates of less than 5 %. To compensate for a low return rate, it is common practice to tag a subsample of 100s to 1000s of sharks or rays (Thorson 1971; Stevens et al. 2000; Kohler and Turner 2001; Kohler et al. 2002). Unfortunately, in dealing with critically endangered species that are found in low numbers, such as P. microdon and G. garricki, this is not always possible. In such circumstances, it is important to maximise the rate of recapture to obtain meaningful sample sizes.
153 Recapture rates are dependent on a number of variables including animal behaviour, mortality, tag retention, habitat (i.e. closed vs. open system), sampling efforts and reporting rates of recaptured animals (Stevens et al. 2000; Kohler and
Turner 2001). Sampling effort in this study was limited to only a few months per year due to the remoteness and extreme nature of the study site environment. Public involvement was therefore relied upon, given that the study area is a popular fishing site for recreational fishers and aboriginal communities. Community education (e.g. permanent signs, informative posters, public and scientific presentations and media outlets), direct involvement of local volunteers and rangers as well as the use of a reward system for reported recaptures were implemented to increase public awareness and involvement in the study. As a result, the majority of recaptures between 2005 and 2010, which occurred >5 days post tagging, were reported by recreational fishers.
However, researchers have been informed by local fishers that numerous recaptures are unreported and tags burned if the animal is consumed, for fear of prosecution. As the reward is unspecified on the tag and consists of a t-shirt, the incentive to report recaptures may not be great enough. Listing high-value rewards for the reporting of tags has been shown to increase the rate of reports (Kohler and Turner 2001; Taylor et al. 2006). Increasing rewards for reports of recaptures or offering entries to a high- reward lottery for each reported recapture should be considered in future work in the
Fitzroy River and King Sound.
Size specific investigations into return rates of P. microdon at liberty for >5 days between 2003 and 2010, found a 9.4 % return rate in P. microdon <1150 mm
TL, but a 17 % return rate in larger (>1780 mm TL) P. microdon. Similarly, monitoring periods from acoustic telemetry (acoustic tags were attached to fin mounted Rototags) were also substantially smaller for YOY (Chapter 4). A smaller
154 return rate of Rototags may be due to increased rates of mortality of YOY P. microdon (Chapter 3). Additionally, the greater growth rate of YOY (Chapter 3) may have resulted in an increased occurrence of premature shedding of Rototags (Stevens et al. 2000; Kohler and Turner 2001). A double tagging study may be useful in assessing the influence of tag retention on recapture rates.
5.4.1.2 Habitat use
The distances between the site of tagging and recapture of conventionally tagged P. microdon, which were at liberty between 5 and 1246 days was negligible, with the exception of a few individuals. Eleven of these fourteen individuals were recaptured at the same location as their initial capture. Thorburn et al. (2007) also reported to have recaptured P. microdon in the same pool in which they were tagged in during a single dry season. It is reasonable to assume this lack of movement between pools in the dry season is a result of restricted up and downstream access brought on by low water levels. However, six of the 11 P. microdon observed not to have moved in this study did have access to neighboring pools during high tides and/or increased stage heights in the wet and/or early dry season (Chapter 4). Having either not of moved, or having returned to the same site between captures suggests some level of habitat fidelity for the pools in which they were recaptured in, at the respective times.
Pristis microdon that were recaptured in the freshwater pools in which they were tagged in, after movement from the site was possible, were occupying pools with upstream physical barriers (e.g. road crossing, Camballin Barrage). Other migrating predatory fishes have been observed to amass and often remain immediately downstream of weirs and dams (Beamesderfer and Rieman 1991; North et al. 1993; Baumgartner 2007). These congregations have been suggested to be the
155 result of the obstruction of upstream movement and/or due to the increased availability of prey (Beamesderfer and Rieman 1991; North et al. 1993; Baumgartner
2007). Recaptured P. microdon may not have moved downstream or may have returned for similar reasons, but further research is needed to properly identify the specific cause.
Observations of P. microdon that were observed to move between pools helped to validate acoustic monitoring results (see Chapter 4). Observations of a
YOY P. microdon moving 10 km upstream (Snag Pool to Langi Crossing) between
June and November 2008, further supported findings of a slow upstream movement of
YOY P. microdon in estuarine pools during the dry season (Chapter 4). The gradual upstream movement of YOY in the estuarine pools into less saline and more thermally stable pools is hypothesised to be a form of behavioural homeostasis
(Chapter 4). In contrast, larger recaptured P. microdon (a 2540 mm TL female and a
~2000 mm TL male) were found to occupy the river mouth area when salinity was at its highest, at sizes when they are thought to transition to marine environments
(Thorburn et al. 2007; see Chapter 4). Simpfendorfer et al. (2011) found larger juvenile smalltooth sawfish (Pristis pectinata) (a marine species that inhabits estuarine rivers as juveniles) to be less sensitive to salinity than smaller conspecifics, possibly because of the decrease in surface area to volume ratio of the fish. Multiple recaptures of the ~2000 mm TL male P. microdon in the river mouth area throughout the year, implies that this individual frequented or remained within this area year round (movement of this P. microdon would have only been possible between King
Sound and estuarine pools of the river in the mid-late dry season [Chapter 4]), regardless of salinity.
156 5.4.2 Satellite tags
5.4.2.1 Utility
The SPOT-5 satellite tags in fin mounts proved to be ineffective for monitoring movements of P. microdon, due to the lack of recorded transmissions. This was unexpected as P. microdon is known to occur in shallow water and has been observed to forage with its dorsal fins exposed (Peverell 2005, 2008; Thorburn 2006; Thorburn et al. 2007). These behaviours should have allowed for numerous successful transmissions. Studies involving P. pectinata (Simpfendorfer and Wiley 2005), the dwarf sawfish (P. clavata) and green sawfish (P. zijsron) (Stevens et al. 2008) have also recently presented similar outcomes from SPOT tags. A minimal number or lack of recorded transmissions from satellite tags have been suggested to be the result of minimal surface time, biofouling, aerial failure and/or premature tag shedding (Hays et al. 2007; Heithaus et al. 2007b; Holdsworth et al. 2009). Simpfendorfer and Wiley
(2005) also found that small (<2000 mm TL) P. pectinata dorsal fins are too flexible to maintain tag aerials at an upright position, which they observed to disrupt transmissions. Dorsal fin structure of P. microdon tagged in this study was rigid enough to keep aerials upright in water and air, and was not thought to have affected transmission success. Mortality may have affected the number of transmissions, but this was thought to be unlikely as tagged P. microdon were in good condition at the time of release. Biofouling was also not thought to have been an issue with transmissions within at least the first month of deployment. Biofouling would have likely only begun to interfere with the conductivity switch after several weeks to months, as observed by Hays et al. (2007), in part due to the high flow conditions in the sound. Additionally, premature shedding was not believed to have been a factor.
The attachment method (i.e. fin mount) used in this study has proven to be of little
157 concern in other elasmobranch research, lasting for at least 30 days and often for hundreds of days (Weng et al. 2005, 2008; Stevens et al. 2010). It can therefore be reasoned that the most probable cause for the paucity of recorded transmissions, in at least the first month, was a lack of surface time. Observations of P. microdon >2000 mm TL to primarily occupy depths >1 m (Chapter 4) support this conclusion.
The SPOT-5 satellite tags were also ineffective in monitoring movements of
G. garricki. Scarcity of transmissions from G. garricki was possibly a result of a lack of surface time, for similar reasons as discussed for P. microdon. Accuracy of the few calculated locations for tag 41103 is uncertain, but the environments of these locations (i.e. turbid nearshore waters within King Sound) were in congruence with habitats where G. garricki has been captured (Chapter 3). Although SPOT-5 tags proved ineffective, successful intermittent transmissions from tag 41103 over several months suggests that G. garricki may occupy surface waters long enough for other satellite tags, such as fast acquiring GPS tags, to effectively track G. garricki.
5.4.2.2 Habitat use
In the absence of location estimates but with success of a small sample of transmissions, times of transmissions were explored to study the behaviour of G. garricki (a lack of data prevented commenting on P. microdon behaviour). A substantially greater number of transmissions from G. garricki tag 41103 at night and during incoming tides suggested this animal to frequent surface waters more during these times. This could indicate that activity of this individual increases at night and/or that it occupies surface waters as it moves in with the tide. During sampling, numerous unidentified sharks were seen ‘finning’ (i.e. swimming at the surface with dorsal fins exposed) with movement into tidal creeks during incoming tides. Contrary
158 to these findings, Pillans et al. (2009) found depth use of three acoustically tracked speartooth sharks (Glyphis glyphis) to not be influenced by time of day or tidal influence in the estuarine reaches of a north Australian river. Conversely, Heupel et al. (2010) did find young C. leucas (a comparable carcharhinid that inhabit similar areas as G. garricki) to occupy more shallow water at night, similar to tag 41103.
Heupel et al. (2010) suggested this behaviour was a result of foraging efforts. As a single sample, no comment can be made about the species as a whole and more detailed observations are needed to confirm depth use patterns of G. garricki.
5.4.3 Conclusion
Results from this study suggest that information derived from conventional tags can be limited in monitoring long-term movements of P. microdon in the confines of a remote river, due to the small number of recaptures. However, the data provided by the few recaptured P. microdon was shown to be useful in complimenting data from acoustic telemetry. As discussed in Chapter 4, acoustic tracking was proven to be valuable in documenting P. microdon movements in riverine environments over several months in the dry season, but not over long periods as tag transmission ended prematurely. As such, it was necessary to pool data from individuals to make inferences regarding long-term movement of P. microdon. Results from mark and recapture methods, being similar to those from acoustic monitoring, helped to validate these inferences.
Mark and recapture data suggested P. microdon to remain or return to pools with physical barriers, even when downstream movement from these pools was possible. This may imply that barriers, like the Camballin Barrage and even simple road crossings, affect the movements of P. microdon. However, it is unclear if
159 individuals are remaining/returning to these pools because their upstream movements are being obstructed, or if P. microdon prefers the habitat created by these barriers.
Due to the limited but informative data provided by mark and recapture methods, it is suggested that conventional tagging be continued. However, it is also recommended that conventional tags be used in association with a more reliable secondary method, like acoustic monitoring. It is also advised that additional efforts need to be undertaken to increase data from mark and recapture methods. As the number of P. microdon tagged is restricted by time available by researchers, enlisting the aid of trained local rangers or fishers could aid in increasing tagging and sampling effort. It is also advisable for future researchers to increase awareness of the tagging program and improve upon the current reward system for reported recaptures to increase the number of these reports. Additionally, due to the rarity of G. garricki captures and ineffectiveness of SPOT-5 tags, other tagging methods should be trialed.
In the opinion of the author, an alternative combination of archival tags, useful in documenting depth use and range of a species, and possibly the use of fast signal acquiring GPS tags may be of better use in monitoring the movements of G. garricki in King Sound.
160 Chapter 6: General discussion
Euryhaline elasmobranchs commonly inhabit rivers and estuaries, using them as nurseries, sanctuaries and foraging areas (Castro 1993; Compagno and Cook 1995;
Martin 2005; Thorburn et al. 2007; Thorburn and Rowland 2008; Pillans et al. 2009).
Dependence of these species on nearshore habitats makes them particularly vulnerable to habitat alteration or loss. However, how or to what extent these fishes are impacted is unclear due to a paucity of information regarding their habitat use behaviours. This is particularly relevant for the critically endangered freshwater sawfish (Pristis microdon) and northern river shark (Glyphis garricki), which inhabit nearshore waters in northern Australia. As the remaining habitats of these species are coming under increasing pressures from human development, including water abstraction, tidal and hydroelectric power as well as mining (Compagno and Cook 1995; Wolanski et al.
2001; Doupé and Pettit 2002; Lindsay and Commander 2005; Thorburn and Morgan
2006), the viability of these species is uncertain. Through the use of multiple tagging techniques, long-term catch data and environmental monitoring, this study generated data on habitat use of juvenile P. microdon in riverine nurseries, and expanded upon the knowledge of the distribution of G. garricki. Using this knowledge, the potential impacts of habitat alterations on P. microdon and G. garricki are discussed below.
6.1 Pristis microdon habitat use
This study demonstrated that juvenile P. microdon inhabit the Fitzroy River for approximately three to five years, where they triple in body length, as also described by Thorburn et al. (2007). Pristis microdon was found to inhabit the entire study site
161 (lower 160 km of the river) throughout the year, although a greater percentage of
YOY were observed to inhabit freshwater, rather than estuarine, pools following relatively large wet seasons. Young of the year and other P. microdon <1650 mm TL that remained in the estuarine pools in the early dry season, appeared to only move upstream away from increasing salinity and diel temperature ranges during the late dry season. At a size of >1850 mm TL, larger individuals were observed to move downstream into the river mouth area during the wet season. Pristis microdon captured in King Sound by this study and by Thorburn et al. (2007), ranged in size between 2100 mm and 2600 mm TL. Together this evidence suggests P. microdon emigrate into the sound starting at a size of around 2000 mm TL. However, males of up to 2400 mm TL and females of up to 2800 mm TL have also been observed within the river (Thorburn et al. 2007), suggesting other factors beyond size/age may influence the timing of P. microdon emigration.
Habitat use of juvenile P. microdon in the Fitzroy River was found to be influenced by several environmental variables, including water depth, flow and light intensity/turbidity. It was also hypothesised that salinity, water temperature and biotic factors, such as prey availability and predation may also influence the habitat use patterns of P. microdon, but this requires further investigation. Habitat use patterns and the influence of environmental change were observed to vary with the size of juvenile P. microdon, likely due to differences in the requirements of the different size classes. As an example, juvenile P. microdon displayed ontogenetic depth partitioning, with YOY inhabiting shallower water, where they likely benefit from decreased risk of predation, and larger individuals inhabiting deeper water, where they likely benefit from increased manoeuvrability. Habitat use behaviours and movements of P. microdon appeared to change over space and time to maximise the
162 overall fitness of individuals in accordance with the status of surrounding abiotic and biotic variables. These changes are likely to aid in the optimisation of net energy gains (e.g. decreasing energy expenditure, increasing foraging efficiency), as well as in decreasing predation and intra-specific competition. Young of the year P. microdon for example, appeared to prefer to move in the direction of tidal flow in the estuarine pools, rather than against it, which would likely reduce energy expenditure.
6.2 Glyphis garricki habitat use
This study demonstrated that young adult male and juvenile G. garricki inhabit turbid and tidal creeks, river mouth areas and nearshore channels throughout the upper King
Sound in the dry season. Additional size classes may inhabit these waters as well, but their capture was likely impeded by size-selectivity of the gillnet mesh sizes that were deployed. Observations of small (900 to 1000 mm TL) juvenile G. garricki within
King Sound (Thorburn and Morgan 2004, Thorburn 2006, Chapter 3) suggests the presence of YOY, which are roughly estimated to be between 500 and 800 mm TL
(Stevens et al. 2005).
Captures of G. garricki were rare. However, G. garricki was relatively abundant in turbid nearshore waters of the upper King Sound when compared to other elasmobranchs captured in these areas (Thorburn 2006; Chapter 3). Conversely, no
G. garricki have been captured or reported to have been captured in areas with low turbididty in King Sound or in freshwaters of surrounding tributaries. Results from this study support previous findings that primary habitat of juvenile G. garricki consist of turbid nearshore waters in close approximation to river mouth areas
(Thorburn and Morgan 2004; Thorburn 2006; Field et al. 2008; Pillans et al. 2009).
163 6.3 Potential impacts of anthropogenic disturbances on habitat use of Pristis microdon and Glyphis garricki
Observing the results from this study, it can be concluded that anthropogenic changes to water depth, flow, turbidity and potentially salinity and/or temperature could impact the habitat use and fitness of juvenile P. microdon, and potentially G. garricki.
Known P. microdon nurseries and G. garricki habitat in northern Australia, have been proposed for, or have been impacted by dams (including hydroelectric), tidal energy barrages, water abstraction programs and mining (Wolanski et al. 2001; Doupé and
Pettit 2002; Lindsay and Commander 2006; Thorburn and Morgan 2006; CSIRO
2009). However, the effects of these developments on P. microdon or G. garricki have not been well studied. Potential impacts of human development on P. microdon and G. garricki movement and habitat use in northern Australia are discussed below, using literature documenting the effects of human development on riverine and estuarine environments in conjunction with the results of this study.
Flow regulating dams constructed for flood control, irrigation and hydropower production have been shown to greatly alter environmental variables including water flow, depth, substrate composition, turbidity and temperature (Banks 1969; Stanford et al. 1996). These alterations have resulted in changes to the composition, seasonal migration, behaviours and survivorship of numerous fishes (Banks 1969; Raymond
1979; Drinkwater and Frank 1994; Gehrke et al. 2002). Based on observations made during this study, the far reaching changes to river flow and depth by dam operations are likely to be the most influential changes to P. microdon movement and habitat use. To control flooding and to store water for future use (i.e. power generation and irrigation) dams often retain river discharge during wet seasons. This diminishes the rise in stage height of rivers and the flushing of fine sediment during the wet season
164 (Young et al. 2004), which reduces river depth. In such conditions, inter-pool movement of P. microdon, which is highly dependent upon river depth, could be reduced. However, unseasonal releases of stored water by dams increase baseflows in the dry season (Doupé and Pettit 2002), which may provide at least YOY P. microdon with greater access between pools throughout the year. Although this may benefit
YOY P. microdon, mistimed peaks of flow in the dry season would create unseasonal decreases in salinity downstream and alter habit use of fishes that inhabit river mouth areas (Drinkwater and Frank 1994), like G. garricki and the dwarf sawfish (Pristis clavata) (Thorburn et al. 2008). These impacts may be reduced through altering spill regimes to simulate natural patterns.
Dams also act as physical barriers, directly influencing upstream and downstream fish movements (North et al. 1993; Lucas et al. 2009). Obstruction by small barrages in rivers can be minor to P. microdon if floods allow movement over or around the barrage in the wet season, when the majority of inter-pool movement occurs. However, the influence of larger dams would be great, as these structures create impassible barriers. Fish ladders, lifts and other dam bypasses can be used to allow fishes access beyond dams (Stuart and Berghuis 2002). However, many of these methods would prove impractical for most juvenile P. microdon, due to their large size.
Like hydroelectric dams, tidal barrages installed in river mouths, bays and tidal creeks, block movement of fish and cause high mortality rates for fish that pass through turbines used to generate power (Dadswell et al. 1986; Davies 1988;
Dadswell and Rulifson 1994). Only when sluices are open is safe movement beyond the barrage possible. Tidal barrages are also known to restrict flow and decrease turbidity (Dadswell et al. 1986; Bryden et al. 2004). Installation of a tidal barrage in
165 tidal creeks of King Sound would foreseeably decrease the access of G. garricki to these areas, which are primary habitat for this species. Proposals were submitted to install such a barrage at the mouth of Doctor’s Creek in King Sound (EPA 1999,
2002), where several G. garricki have been captured (Thorburn and Morgan 2004;
Thorburn 2006; Chapter 3). As one of the King Sound’s largest mangrove rich tidal creeks, also likely rich in prey (Bell et al. 1984; Robertson and Duke 1987), loss of such a system would negatively impact G. garricki. This impact may not be critical in King Sound due to the abundance of tidal creeks and rivers in the area. However, loss of such an area in a smaller system would likely have a much greater effect on G. garricki. Alternative methods of tidal energy, such as tidal turbines, which do not restrict fish or water movement, would potentially have a lower impact on G. garricki, although noise generation and/or damage from the turbines may still pose a threat
(Pelc and Fujita 2002).
Water abstraction and diversion used for potable supply, irrigation and mining decreases river baseflow and depth (Kingsford et al. 1998; Thomas and Sheldon
2000). A decrease in water depth from water abstraction will decrease the time in which P. microdon can move between pools. In years of increased precipitation, the impact of a drawdown of water would likely be minimal on P. microdon movement.
However, in years with a minor wet season, abstraction or diversion of water would be compounded by naturally low water levels. This would severely limit P. microdon movement between pools and potentially increase intra-specific competition.
It is evident from the results of this study, that the discussed developments could alter the fitness of juvenile P. microdon and G. garricki if managed poorly.
Conservation of these species would be maximised by denying future development of important P. microdon and G. garricki habitat. However, the increasing needs of
166 growing human populations make such a feat unrealistic. It is therefore important that future development of known P. microdon nurseries and G. garricki habitat be well regulated to minimise the impact of these disturbances. This can be accomplished in part by:
1) minimising flow regulation, and mimicking natural flow patterns
2) reducing water abstraction and flow control in years with relatively little
precipitation
3) installing less obstructive tidal power generation methods in place of tidal
barrages
Additionally, it is suggested that future risk assessment of P. microdon consider the differences in habitat use of the various life stages (i.e. YOY, juvenile, adult).
6.4 Future research
This study has advanced the understanding of the habitat use behaviours and/or distribution of P. microdon and G. garricki through the documentation of their habitat, demography, mortality, growth and movements. However, further research of
P. microdon and G. garricki is needed to better understand the habitat use behaviours of these species. It is recommended that future studies target the following to further advance our understanding of these endangered species:
Catch per unit effort (CPUE) of P. microdon varied seasonally and annually,
which made it difficult to ascertain overall long-term trends. Monitoring of
seasonal CPUE should be continued with this population to better understand
167 a ‘normal’ baseline CPUE, and to be able to detect future changes. It is also
recommended that future research in the Fitzroy River include additional
sample sites between the estuarine and freshwater pools sampled, to obtain a
more accurate representation of P. microdon CPUE in the river as a whole.
Wet season sampling and tagging should be included in future studies to
investigate the potential of seasonal growth and to explore large-scale
upstream movements. Further exploration of upstream movement would help
to explain the relationship between wet season size and parturition date with
upstream movement of YOY P. microdon.
Passive acoustic monitoring in the Fitzroy River should be continued, with
greater emphasis in freshwater pools. Continuation of this method would
increase sample sizes and increase the power of analyses. As no information
of depth use of YOY P. microdon in the freshwater pools was documented,
tagging YOY in this region would allow for a better understanding of
ontogenetic habitat partitioning. In addition, it is advisable to implant acoustic
transmitters within at least YOY to increase tag retention times.
Observing the poor performance of SPOT-5 satellite tags on P. microdon and
G. garricki, and the poor performance of passive acoustic monitoring in the
mouth of the Fitzroy River, it is advisable to conduct preliminary tests of other
electronic tagging methods to monitor the movements of these species. Other
potential methods that may prove of use include: active acoustic tracking, fast
acquiring GPS tags and/or archival tags.
168 To ensure the long-term survival and recovery of the world’s relatively few remaining P. microdon and G. garricki populations, it is important to educate and involve local communities and fishers with conservation and management efforts, increase enforcement of current regulations and conserve the natural balance of vital areas, such as the Fitzroy River and King Sound.
169 References
Abril, G., Etcheber, H., Delille, B., Frankignolle, M. and Borges, A. V. (2003). Carbonate dissolution in the turbid and eutrophic Loire estuary. Marine Ecology Progress Series 259, 129-138.
Ackerman, J. T., Kondratieff, M. C., Matern, S. A. and Cech, J. J. (2000). Tidal influence on spatial dynamics of leopard sharks, Triakis semifasciata, in Tomales Bay, California. Environmental Biology of Fishes 58, 33-43.
Agassiz, L. (1843). ‘Recherches sur les poissons fossiles’. Vol. 3. (Contenant l’Histoire de l’Ordre des Placoides: Neuchâtel, Switzerland).
Anderson, J. T. and Rose, G. A. (2001). Offshore spawning and year-class strength of northern cod (2J3KL) during the fishing moratorium, 1994-1996. Canadian Journal of Fisheries and Aquatic Sciences 58, 1386-1394.
Arnold, G. P. and Dewar, H. (2001). Electronic tags in marine fisheries research: a 30 year perspective. In ‘Electronic Tagging and Tracking in Marine Fisheries’. (Eds J. R. Sibert and J. L. Nielsen). pp. 7-64. (Kluwer Academic Publishers: The Netherlands).
Arscott, D. B., Tockner, K. and Ward, J. V. (2001). Thermal heterogeneity along a braided floodplain river (Tagliamento River, northeastern Italy). Canadian Journal of Fisheries and Aquatic Sciences 58, 2359-2373.
The Australian Government Department of the Environment, Water, Heritage and the Arts (2009). ‘Australian Natural Resources Atlas: Water resources – Overview –Western Australia’. Available at http://www.anra.gov.au/topics/water/management/wa/basin-fitzroy-river- wa.html [accessed 22 April 2010].
Balls, P. W., Brockie, N., Dobson, J. and Johnston, W. (1996). Dissolved oxygen and nitrification in the upper Forth Estuary during summer (1982-1992): patterns and trends. Estuarine, Coastal and Shelf Science 42, 117-134.
Banks, J. W. (1969). A review of the literature on the upstream migration of adult salmonids. Journal of Fish Biology 1, 85-136.
Basson, M., Rosenberg, A. A. and Beddington, J. R. (1988). The accuracy and reliability of two new methods for estimating growth parameters from length- frequency data. ICES Journal of Marine Science 44, 277-285.
Baumgartner, L. J. (2007). Diet and feeding habits of predatory fishes upstream and downstream of a low-level weir. Journal of Fish Biology 70, 879-894.
170 Beamesderfer, R. C. and Rieman, B. E. (1991). Abundance and distribution of northern squawfish, walleyes, and smallmouth bass in John Day Reservoir, Columbia River. Transactions of the American Fisheries Society 120, 439- 447.
Beck, M. W., Heck, K. L. Jr, Able, K. W., Childers, D. L., Eggleston, D. B., Gillanders, B. M., Halpern, B., Hays, C. G., Hoshino, K., Minello, T. J., Orth, R. J., Sheridan, P. F. and Weinstein, M. P. (2001). The identification, conservation, and management of estuarine and marine nurseries for fish and invertebrates. Bioscience 51, 633-641.
Bell, J. D., Pollard, D. A., Burchmore, J. J., Pease, B. C. and Middleton, M. J. (1984). Structure of a fish community in a temperate tidal mangrove creek in Botany Bay, New South Wales. Australian Journal of Marine and Freshwater Research 35, 33-46.
Begg, G. A. (2005). Life history parameters. In ‘Stock identification methods’. (Eds S. X. Cadrin, K. D. Friedland and J. R. Waldman.) pp. 119-150. (Elsevier Academic Press: New York).
Bethea, D. M., Buckel, J. A. and Carlson, J. K. (2004). Foraging ecology of the early life stages of four sympatric shark species. Marine Ecology Progress Series 268, 245-264.
Bhargava, D. S. and Mariam, D. (1991). Effects of suspended particle size and concentration on reflectance measurements. Photogrammetric Engineering and Remote Sensing 57, 519-529.
Boynton, W. R. and Kemp, W. M. (1985). Nutrient regeneration and oxygen consumption by sediments along an estuarine salinity gradient. Marine Ecology Progress Series 23, 45-55.
Bruce, B. D., Stevens, J. D. and Malcolm, H. (2006). Movements and swimming behaviour of white sharks (Carchardon carcharius) in Australian waters. Marine Biology 150, 161-172.
Bryden, I. G., Grinstead, T. and Melville, G. T. (2004). Assessing the potential of a simple tidal channel to deliver useful energy. Applied Ocean Research 26, 198-204.
Burns, J. M. and Castellini, M. A. (1998). Dive data from satellite tags and time-depth recorders: a comparison in Weddell Seal pups. Marine Mammal Science 14, 750-764.
Carlisle, A. B. and Starr, R. M. (2009). Habitat use, residency, and seasonal distribution of female leopard sharks Triakis semifasciata in Elkhorn Slough, California. Marine Ecology Progress Series 380, 213-228.
171 Carlson, J. K., Heupel, M. R., Bethea, D. M. and Hollensead, L. D. (2008). Coastal habitat use and residency of juvenile Atlantic sharpnose sharks (Rhizoprionodon terraenovae). Estuaries and Coasts 31, 931-940.
Caron, F., Hatin, D. and Fortin, R. (2002). Movements and aggregation areas of adult Atlantic sturgeon (Acipenser oxyrinchus) in the St Lawrence River estuary, Québec, Canada. Journal of Applied Ichthyology 18, 586-594.
Carrier, J. C. and Luer, C. A. (1990). Growth rates in the Nurse Shark, Ginglymostoma cirratum. Copeia 1990, 686-692.
Cartamil, D. P., Vaudo, J. J., Lowe, C. G., Wetherbee, B. M. and Holland, K. N. (2003). Diel movement patterns of the Hawaiian stingray, Dasyatis lata: implications for ecological interactions between sympatric elasmobranch species. Marine Biology 142, 841-847.
Castro, J. I. (1993). The shark nursery of Bulls Bay, South Carolina, with a review of the shark nurseries of the southeastern coast of the United States. Environmental Biology of Fishes 38, 37-48.
Chalfoun, A. D., Martin, T. E. (2007). Assessments of habitat preferences and quality depend on spatial scale and metrics of fitness. Journal of Applied Ecology 44, 983-992.
Chaudhury, R. R., Sobrinho, J. A. H., Wright, R. M. and Screenivas, M. (1998). Dissolved oxygen modeling of the Blackstone River (Northeastern United States). Water Research 32, 2400-2412.
Chen, S. and Watanabe, S. (1989). Age dependence of natural mortality coefficient in fish population dynamics. Nippon Suisan Gakkaishi 55, 205-208.
Chidlow, J. A. (2007). First record of the freshwater sawfish, Pristis microdon, from southwestern Australian waters. Records of the Western Australian Museum 23, 307-308.
Chiu, S. and Abrahams, M. V. (2010). Effects of turbidity and risk of predation on habitat selection decisions by fathead minnow (Pimephales promelas). Environmental Biology of Fishes 87, 309-316.
Clements, S., Jepsen, D., Karnowski, M. and Schreck, C. B. (2005). Optimization of an acoustic telemetry array for detecting transmitter-implanted fish. North American Journal of Fisheries Management 25, 429-436.
Collins, M. A., Bailey, D. M., Ruxton, G. D. and Priede, I. G. (2005). Trends in body size across an environmental gradient: a differential response in scavenging and non-scavenging demersal deep-sea fish. Proceedings of the Royal Society B 272, 2051-2057.
172 Collins, A. B., Heupel, M. R. and Motta, P. J. (2007). Residence and movement patterns of cownose rays Rhioptera Bonasus within a south-west Florida estuary. Journal of Fish Biology 71, 1159-1178.
Commonwealth of Australia, Bureau of Meteorology (2010). ‘Climate statistics for Australian locations’. Available at http://www.bom.gov.au/climate/averages/tables/cw_003032.shtml [accessed 1 March 2010].
Compagno, L. J. V. (1984). Family Carcharhinidae – Requiem sharks. In ‘FAO Species Catalogue. Volume 4, Part 2: Sharks of the World’ (Ed L. J. V. Compagno). pp. 445-501. (Food and Agriculture Organization of the United Nations: Rome).
Compagano, L. J. V. (1990). Alternative life-history styles of cartilaginous fishes in time and space. Environmental Biology of Fishes 28, 33-75.
Compagno, L. J. V. and Cook, S. F. (1995). The exploitation and conservation of freshwater elasmobranchs: status of taxa and prospects for the future. Journal of Aquariculture and Aquatic Sciences 8, 62-90.
Compagno, L. J. V., Cook, S. F. and Fowler, S. L. (2006). The IUCN Red List of Threatened Species: Pristis microdon (Version 2009.2)’. Available at http://www.iucnredlist.org [accessed 17 January 2010].
Compagno, L. J. V. and Last, P. R. (1998). Order Pristiformes, Pristidae - Sawfishes. In ‘FAO Species Identification Guide for Fishery Purposes. The Living Marine Resources of the Western Central Pacific. Volume 3: Batoid Fishes, Chimaeras and Bony Fishes Part 1 (Elopidae to Linophrynidae)’. (Eds K. E. Carpenter and V. H. Niem). pp. 1410-1417. (FAO: Rome).
Compagno, L. J. V. and Niem, V. H. (1998). Carcharhinidae - Requiem sharks. In ‘FAO Species Identification Guide for Fisheries Purposes. The Living Marine Resources of the Western Central Pacific. Volume 2: Cephalopods, crustaceans, holothurians and sharks’. (Eds K. E. Carpenter and V. H. Niem). pp. 1312-1360. (FAO: Rome).
Compagno, L. J. V., White, W. T. and Last, P. R. (2008). Glyphis garricki sp. nov., a new species of river shark (Carcharhiniformes: Carcharhinidae) from northern Australia and Papua New Guinea, with a redescription of Glyphis glyphis (Müller and Henle, 1839). In ‘Descriptions of New Australian Chondrichthyans’. (Eds P. R. Last, W. T. White and J. J. Pogonoski). pp. 203- 226. (CSIRO Marine and Atmospheric Research: Hobart).
Conover, W. J. and Iman, R. L. (1981). Rank transformations as a bridge between parametric and nonparametric statistics. The American Statistician 35, 124- 129.
173 Cook, S. F., Compagno, L. J. V. and Last, P. R. (2006). ‘The IUCN Red List of Threatened Species: Pristis clavata (Version 2011.1)’. Available at http://www.iucnredlist.org [accessed on 22 July 2011].
Cortés, E. (1998). Demographic analysis as an aid in shark stock assessment and management. Fisheries Research 39, 199-208.
Cortés, E. (2002). Incorporating uncertainty into demographic modeling: application to shark populations and their conservation. Conservation Biology 16, 1048- 1062.
Cortés, E. and Parson, G. R. (1996). Comparative demography of two populations of the bonnethead shark (Sphyrna tiburo). Canadian Journal of Fisheries and Aquatic Sciences 53, 709-718.
Cowan, J. L. W., Pennock, J. R. and Boynton, W. R. (1996). Seasonal and interannual patterns of sediment-water nutrient and oxygen fluxes in Mobile Bay, Alabama (USA): regulating factors and ecological significance. Marine Ecology Progress Series 141, 229-245.
CSIRO (2009). Water in the Fitzroy region. In ‘Water in the Timor Sea Drainage Division’. pp. 61-128. Report to the Australian Government, Canberra.
Cyrus, D. P. and Blaber, S. J. M. (1992). Turbidity and salinity in a tropical northern Australian estuary and their influence on fish distribution. Estuarine and Coastal Shelf Science 35, 545-563.
Dadswell, M. J. and Rulifson, R. A. (1994). Macrotidal estuaries: a region of collision between migratory marine animals and tidal power development. Biological Journal of the Linnean Society 51, 93-113.
Dadswell, M. J., Rulifson, R. A. and Daborn, G. R. (1986). Potential impact of large- scale tidal power developments in the upper Bay of Fundy on fisheries resources of the Northwest Atlantic fisheries. Fisheries 11, 26-35.
Dai, M., Guo, X., Zhai, W., Yuan L., Wang, B., Wang, L., Cai, P., Tang, T. and Cai, W. (2006). Oxygen depletion in the upper reach of the Pearl River estuary during a winter drought. Marine Chemistry 102, 159-169.
Davies, J. K. (1988). A review of information relating to fish passage through turbines: implications of tidal power schemes. Journal of Fish Biology 33, 111-126.
Department of Sustainability, Environment, Water, Population and Communities (2011a). ‘Species Profile and Threats Database: Pristis microdon’. Available at http://www.environment.gov.au/sprat [accessed on 22 July 2011].
Department of Sustainability, Environment, Water, Population and Communities (2011b). ‘Species Profile and Threats Database: Glyphis garricki’. Available at http://www.environment.gov.au/sprat [accessed on 22 July 2011].
174 Dewar, H., Mous, P., Domeier, M., Muljadi, A., Pet, J. and Whitty, J. (2008). Movements and site fidelity of the giant manta ray, Manta birostris, in the Komodo Marine Park, Indonesia. Marine Biology 155, 121-133.
Doupé, R. G. and Pettit, N. E. (2002). Ecological perspectives on regulation and water allocation for the Ord River, Western Australia. River Research and Applications 18, 307-320.
Drinkwater, K. F. and Frank, K. T. (1994). Effects of river regulation and diversion on marine fish and invertebrates. Aquatic Conservation: Marine and Freshwater Ecosystems 4, 135-151.
Dulvy, N. K. and Forrest, R. E. (2010). Life histories, population dynamics, and extinction risks in chondrichthyans. In ‘Sharks and Their Relatives II: Biodiversity, Adaptive Physiology, and Conservation’. (Eds J. C. Carrier, J. A. Musick and M. R. Heithaus). pp. 639-679. (Taylor and Francis Group: Boca Raton, Florida).
Duncan, K. M. and Holland, K. N. (2006). Habitat use, growth rates and dispersal patterns of juvenile scalloped hammerhead sharks (Sphyrna lewini) in a nursery habitat. Marine Ecology Progress Series 312, 211-221.
Ebert, D. A. (2002). Ontogenetic changes in the diet of the sevengill shark (Notorynchus cepedianus). Marine and Freshwater Research 53, 517-523.
Ebner, B. C., Johnston, L. and Lintermans, M. (2009). Radio-tagging and tracking of translocated cod (Maccullochella macquariensis: Percichthyidae) in an upland river. Marine and Freshwater Research 60, 346-355.
Eckert, S. A. and Stewart, B. S. (2001). Telemetry and satellite tracking of whale sharks, Rhincodon typus, in the Sea of Cortez, Mexico, and the North Pacific Ocean. Environmental Biology of Fishes 60, 299-308.
Environmental Protection Authority (1999). Derby Tidal Power Project: Derby Hydro Power Pty Ltd. Bulletin 942, Perth.
Environmental Protection Authority (2002). Derby Tidal Power Project: Tidal Energy Australia. Bulletin 1071, Perth.
Evans, E. C., McGregor, G. R. and Petts, G. E. (1994). River energy budgets with special reference to river bed processes. Hydrological Processes 12, 575-595.
Faria, V. V. (2007). Taxonomic review, phylogeny, and geographical population structure of the sawfishes (Chondrichthyes, Pristiformes). PhD Thesis, Iowa State University.
175 Field, I. C., Carters, R., Buckworth, R. C., Meekan, M. G. and Bradshaw, C. J. A. (2008). Distribution and abundance of Glyhpis and sawfishes in northern Australia and their potential interactions with commercial fisheries. Final report to the Department of the Environment, Water, Heritage and the Arts, Australian Government, Canberra.
Fischhoff, I. R., Sundaresan, S. R., Cordingley, J. and Rubenstein, D. I. (2007). Habitat use and movements of plains zebra (Equus burchelli) in response to predation danger from lions. Behavioural Ecology 18, 725-729.
Foster, I. D. L., Millington, R. and Grew, R. G. (1992). The impact of particle size controls on stream turbidity measurement; some implications for suspended sediment yield estimation. In ‘Erosion and Sediment Transport Monitoring Programmes in River Basins (Proceedings of the Oslo Symposium, August 1992)’. (Eds J. Bogen, D. E. Walling and T. J. Day). pp. 51-62. (International Association of Hydrological Sciences Publications: Oxfordshire, United Kingdom).
Francis, M.P. and Duffy, C. (2002). Distribution, seasonal abundance and bycatch of baskig sharks (Cetorhinus maximus) in New Zealand, with observations on their winter habitat. Marine Biology 140, 831-842.
Fritz, H., Said, S. and Weimerskirch, H. (2003). Scale-dependent hierarchical adjustments of movement patterns in a long-range foraging seabird. Proceeds of the Royal Society London 270, 1143-1148.
Fuentes-Yaco, C., de León, D. A. S., Monreal-Gómez, M. A. and Vera-Herrera, F. (2001). Environmental forcing in a tropical estuarine ecosystem: the Palizada River in the southern Gulf of Mexico. Marine and Freshwater Research 52, 735-744.
García-Charton, J. A., Pérez-Ruzafa, Á., Sánchez-Jerez, P., Bayle-Sempere, J. T., Reñones, O. and Moreno, D. (2004). Multi-scale spatial heterogeneity, habitat structure, and the effect of marine reserves on Western Mediterranean rocky reef fish assemblages. Marine Biology 144, 161-182.
Gehrke, P. C., Gilligan, D. M. and Barwick, M. (2002). Changes in fish communities of the Shoalhaven River 20 years after construction of Tallowa Dam, Australia. River Research and Applications 18, 265-286.
Gibson, R. N. (1997). Behaviour and the distribution of flatfishes. Journal of Sea Research 37, 241-256.
Grubbs, R. D. (2010). Ontogenetic shifts in movements and habitat use. In ‘Sharks and Their Relatives II: Biodiversity, Adaptive Physiology, and Conservation’. (Eds J. C. Carrier, J. A. Musick and M. R. Heithaus). pp. 319-350. (Taylor and Francis Group: Boca Raton, Florida).
176 Gruber, S. H., de Marignac, J. R. C. and Hoenig, J. M. (2001). Survival of juvenile lemon sharks at Bimini, Bahamas, estimated by mark-depletion experiments. Transactions of the American Fisheries Society 130, 376-384.
Gruber, S. H., Nelson, D. R. and Morrissey, J. F. (1998). Patterns of activity and space utilization of lemon sharks, Negaprion brevirostris, in a shallow Bahamian Lagoon. Bulletin of Marine Science 43, 61-76.
Hall, L. S., Krausman, P. R. and Morrison, M. L. (1997). The habitat concept and a plea for standard terminology. Wildlife Society Bulletin 25, 173-182.
Hardiman, J. M., Johnson, B. M. and Martinez, P. J. (2004). Do predators influence the distribution of age-0 kokanee in a Colorado reservoir? Transactions of the American Fisheries Society 133, 1366-1378.
Hays, G. C., Bradshaw, C. J. A., James, M. C., Lovell, P. and Sims, D. W. (2007). Why do Argos Satellite tags deployed on marine animals stop transmitting? Journal of Experimental Marine Biology and Ecology 349, 52-60.
Heithaus, M. R. (2001). The biology of tiger sharks, Galeocerdo cuvier, in Shark Bay, Western Australia: sex ratio, size distribution, diet, and seasonal changes in catch rates. Environmental Biology of Fishes 61, 25-36.
Heithaus, M. R. and Dill, L. M. (2006). Does tiger shark predation risk influence foraging habitat use by bottlenose dolphins at multiple spatial scales? Oikos 114, 257-264.
Heithaus, M. R., Frid, A., Wirsing, A. J., Dill, L. M., Fourqurean, J. W., Burkholder, D., Thomson, J. and Bejder, L. (2007a). State-dependent risk-taking by green sea turtles mediates top-down effects of tiger shark intimidation in a marine ecosystem. Journal of Animal Ecology 76, 837-844.
Heithaus, M. R., Hamilton, I. M., Wirsing, A. J. and Dill, L. M. (2006). Validation of a randomization procedure to assess animal habitat preferences: microhabitat use of tiger sharks in a seagrass ecosystem. Journal of Animal Ecology 75, 666-676.
Heithaus, M. R., Wirsing, A. J., Dill, L. M. and Heithaus, L. I. (2007b). Long-term movements of tiger sharks satellite-tagged in Shark Bay, Western Australia. Marine Biology 151, 1455-1461.
Henningsen, A. D. (1994). Tonic immobility in 12 elasmobranchs: use as an aid in captive husbandry. Zoo Biology 13, 325-332.
Heupel, M. R. (2007). Exiting Terra Ceia Bay: examination of cues stimulating migration from a summer nursery area. In ‘Shark Nursery Grounds of the Gulf of Mexico and the East Coast Waters of the United States’. (Eds C. T. McCandless, N. E. Kohler and H. L. Jr. Pratt). pp. 265-280. (American Fisheries Society: Bethesda).
177 Heupel, M. R., Carlson, J. K. and Simpfendorfer, C. A. (2007). Shark nursery areas: concepts, definition, characterization and assumptions. Marine Ecology Progress Series 337, 287-297.
Heupel, M. R., Semmens, J. M. and Hobday, A. J. (2006). Automated acoustic tracking of aquatic animals: scales, design and deployment of listening station arrays. Marine and Freshwater Research 57, 1-13.
Heupel, M. R. and Simpfendorfer, C. A. (2002). Estimation of mortality of juvenile blacktip sharks, Carcharhinus limbatus, within a nursery area using telemetry data. Canadian Journal of Fisheries and Aquatic Sciences 59, 624-632.
Heupel, M. R. and Simpfendorfer, C. A. (2008). Movement and distribution of young bull sharks Carcharhinus leucas in a variable estuarine environment. Aquatic Biology 1, 277-289.
Heupel, M. R., Simpfendorfer, C. A. and Bennet, M. B. (1998). Analysis of tissue responses to fin tagging in Australian carcharhinids. Journal of Fish Biology 52, 610-620.
Heupel, M. R., Simpfendorfer, C. A. and Hueter, R. E. (2003). Running before the storm: blacktip sharks respond to falling barometric pressure associated with Tropical Storm Gabrielle. Journal of Fish Biology 63, 1357-1363.
Heupel, M. R., Simpfendorfer, C. A. and Hueter, R. E. (2004). Estimation of shark home ranges using passive monitoring techniques. Environmental Biology of Fishes 71, 135-142.
Heupel, M. R., Yeiser, B. G., Collins, A. B., Ortega, L. and Simpfendorfer, C. A. (2010). Long-term presence and movement patterns of juvenile bull sharks, Carcharhinus leucas, in an estuarine river system. Marine and Freshwater Research 61, 1-10.
Hight, B. V. and Lowe, C. G. (2007). Elevated body temperatures of adult female leopard sharks, Triakis semifasciata, while aggregating in shallow nearshore embayments: evidence for behavioral thermoregulation? Journal of Experimental Marine Biology and Ecology 352, 114-128.
Hoenig, N. and Chaudary Hanumara, R. (1982). A statistical study of a seasonal growth model for fishes. Technical Report to the Department of Computer Sciences and Statistics, University of Rhode Island.
Holdsworth, J. C., Sippel, T. J. and Block, B. A. (2009). Near real time satellite tracking of striped marlin (Kajikia audax) movements in the Pacific Ocean. Marine Biology 156, 505-514.
Holland, K. N., Wetherbee, B. M., Lowe, C. G. and Meyer, C. G. (1999). Movements of tiger sharks (Galeocerdo cuvier) in coastal Hawaiian waters. Marine Biology 134, 665-673.
178 Holland, K. N., Wetherbee, B. M., Peterson, J. D. and Lowe, C. G. (1993). Movements and distribution of hammerhead shark pups on their natal grounds. Copeia 1993, 495-502.
Hopkins, T. E. and Cech, J. J. Jr. (2003). The influence of environmental variables on the distribution and abundance of three elasmobranchs in Tomales Bay, California. Environmental Biology of Fishes 60,279-291.
Hugie, D. M. and Dill, L. M. (1994). Fish and game: a game theoretic approach to habitat selection by predators and prey. Journal of Fish Biology 45, 151-169.
Hulbert, L. B., Aires-da-Silva, A. M., Gallucci, V. F. and Rice, J. S. (2006). Seasonal foraging movements and migratory patterns of female Lamna ditropis tagged in Prince William Sound, Alaska. Journal of Fish Biology 67, 490-509.
Hydstra database - time-series data (2009). Department of Water, Water Information Provision section, Perth.
Imbrock, F., Appenzeller, A. and Eckmann, R. (1996). Diel and seasonal distribution of perch in Lake Constance: a hydroacoustic study in situ observations. Journal of Fish Biology 49, 1-13.
Ishihara, H., Taniuchi, T. and M. Shimizu, M. (1991). Sexual dimorphism in number of rostral teeth in the sawfish, Pristis microdon collected from Australia and Papua New Guinea. The University Museum, University of Tokyo, Nature and Culture 3, 83-89.
IUCN. (2009). ‘IUCN Red List of Threatened Species (Version 2009.2)’. Available at www.iucnredlist.org [accessed 5 December 2009].
Jellyman, D. (2009). A review of radio and acoustic telemetry studies of freshwater fish in New Zealand. Marine and Freshwater Research 60, 321-327.
Jennings, J. N. (1975). Desert dunes and estuarine fill in the Fitzroy estuary (north- western Australia). Catena 2, 215-262.
Jennings, J. N. and Coventry, R. J. (1973). Structure and texture of a gravelly barrier island in the Fitzroy estuary, Western Australia, and the role of mangroves in the shore dynamics. Marine Geology 15, 145-167.
Jennings, D. E., Gruber, S. H., Franks, B. R., Kessel, S. T. and Robertson, A. L. (2008). Effects of large-scale anthropogenic development on juvenile lemon shark (Negaprion brevirostris) populations of Bimini, Bahamas. Environmental Biology of Fishes 83, 369-377.
Jensen, A. L. (1996). Beverton and Holt life history invariants result from optimal trade-off of reproduction and survival. Canadian Journal of Fisheries and Aquatic Sciences 53, 820-822.
179 Jerlov, N. G. (1968). ‘Optical Oceanography’. (Elsevier Publishing Company: Amsterdam).
Jonsson, N., Jonsson, B. and Hansen, P. (1990). Partial segregation in the timing of migration of Atlantic salmon of different ages. Animal Behaviour 40, 313-321.
Jorgensen, S. J., Reeb, C. A., Chapple, T. K., Anderson, S., Perle, C., Sommeran, S. R. V., Frtiz-Cope, C., Brown, A. C., Klimley, A. P. and Block, B.A. (2009). Philopatry and migration of Pacific white sharks. Proceedings of the Royal Society 277, 679-688.
Junk, W. J., Bayley, P. B. and Sparks, R. E. (1989). The flood pulse concept in river- floodplain system. In ‘Proceedings of the International Large River Symposium, Florence, 12-17 June 1989’. (Ed. D. P. Dodge). pp. 110-127. (Canadian Special Publication of Fisheries and Aquatic Sciences: 106).
Kemp, W. M. and Boynton, W. R. (1980). Influence of biological and physical processes on dissolved oxygen dynamics in an estuarine system: implications for the measurement of community metabolism. Estuarine and Coastal Marine Science 11, 407-431.
Kemp, W. M., Sampou, P. A., Garber, J., Tuttle, J. and Boynton, W. R. (1992). Seasonal depletion of oxygen from bottom waters of Chesapeake Bay: roles of benthic and planktonic respiration and physical exchange processes. Marine Ecology Progress Series 85, 137-152.
Kerford, M. R., Wirsing, A. J., Heithaus, M. R. and Dill, L. M. (2008). Danger on the rise: diurnal tidal state mediates an exchange of food for safety by the bar- bellied sea snake Hydrophis elegans. Marine Ecology Progress Series 358, 289-294.
Kinney, M. J. and Simpfendorfer, C. A. (2008). Reassessing the value of nursery areas to shark conservation and management. Conservation Letters 2, 53-60.
Kirk, J. T. O. (1985). Effects of suspensoids (turbidity) on penetration of solar radiation in aquatic ecosystems. Hydrobiologia 125, 195-108.
Klimley, A. P., Burney, J. L. B., Cantara, K. M., Richert, J. E., Davis, S. F. and Sommeran, S. V. (2001). Radio-acoustic positioning as a tool for studying site-specific behavior of the white shark and other large marine species. Marine Biology 138, 429-446.
Klimley, A. P., Voegeli, F., Beavers, S. C. and Le Boeuf, B. J. (1998). Automated listening stations for tagged marine fishes. Marine Technology Society Journal 32, 94-101.
Kingsford, R. T., Boulton, A. J. and Puckridge, J. T. (1998). Challenges in managing dryland rivers crossing political boundaries: lessons from Cooper Creek and the Paroo River central Australia. Aquatic Conservation: Marine and Freshwater Ecosystems 8, 361-378.
180 Kirkwood, G. P., Aukland, R. and Zara, S. J. (2001). Length Frequency Distribution Analysis (LFDA) (Version 5.0). (MRAG Ltd: London, UK).
Kohler, N. E. and Turner, P. A. (2001). Shark tagging: a review of conventional methods and studies. Environmental Biology of Fishes 60, 191-223.
Kohler, N. E., Turner, P. A., Hoey, J. J., Natanson, L. J. and Briggs, R. (2002). Tag and recapture data for three pelagic shark species: blue shark (Prionace glauca), shortfin mako (Isurus xyrinchus), and porbeagle (Lamna nasus) in the North Atlantic Ocean. Collective Volume of Scientific Papers of the International Commission for the Conservation of Atlantic Tunas 54, 1231- 1260.
Krause, J., Staaks, G. and Mehner, T. (1998). Habitat choice in shoals of roach as a function of water temperature and feeding rate. Journal of Fish Biology 53, 377-386.
Lacroix, G. L., Knox, D. and Stokesbury, M. J. W. (2005). Survival and behaviour of post-smolt Atlantic salmon in coastal habitat with extreme tides. Journal of Fish Biology 66, 485-498.
Lambeck, R. J. (1997). Focal species: a multi-species umbrella for nature conservation. Conservation Biology 11, 849-856.
Langan, S. J., Johnston, L, Donaghy, M. J., Youngson, A. F., Hay, D. W. and Soulsby, C. (2001). Variation in river water temperatures in an upland stream over a 30-year period. The Science of the Total Environment 265, 195-207.
Larson, H. K. (2000). Report to Parks Australia on estuarine fish monitoring of Kakadu National Park, northern Australia, Australia. Museum and Art Gallery of the Northern Territory, Darwin.
Larson, L. L. and Larson, S. L. (1996). Riparian shade and stream temperature: a perspective. Rangelands 18, 149-152.
Last, P. R. and Stevens, J. D. (1994). ‘Sharks and Rays of Australia’. (CSIRO Publishing: Melbourne).
Last, P. R. and Stevens, J. D. (2009). ‘Sharks and Rays of Australia’. 2nd Edn. (CSIRO Publishing: Melbourne).
Latham, J. (1794). An essay on the various species of sawfish. Transaction of the Linnean Society of London 2, 273-282.
Li, D. J., Zhang, J., Huang, D. J., Wu, Y. and Liang, J. (2002). Oxygen depletion of the Changjiang (Yangtze River) Estuary. Science in China Series D-Earth Sciences 45, 1137-1146.
181 Lindsay, R. P. and Commander D. P. (2006). Hydrogeological assessment of the Fitzroy alluvium. Report to Western Australia Department of Environment and Conservation, Perth.
Lorenzen, K. (1996). The relationship between body weight and natural mortality in juvenile and adult fish: a comparison of natural ecosystems and aquaculture. Journal of Fish Biology 49, 627-647.
Lucas, M. C., Bubb, D. H., Jang, M., Ha, K. and Masters, J. E. G. (2009). Availability of and access to critical habitats in regulated rivers: effects of low-head barriers on threatened lampreys. Freshwater Biology 54, 621-634.
Lucifora, L. O., Menni, R. C. and Escalante, A. H. (2005). Reproduction, abundance and feeding habits of the broadnose sevengill shark Notorynchus cepedianus in north Patagonia, Argentina. Marine Ecology Progress Series 289, 237-244.
MacArthur, R. and Levins, R. (1964). Competition, habitat selection, and character displacement in a patchy environment. Zoology 51, 1207-1210.
MacArthur, R. H. and Pianka, E. R. (1966). On optimal use of a patchy environment. The American Naturalist 100, 603-609.
Malcolm, I. A., Hannah, D. M., Donaghy, M. J., Soulsby, C. and Youngson, A. F. (2004). The influence of riparian woodland on the spatial and temporal variability of stream water temperatures in an upland salmon stream. Hydrology and Earth System Sciences 8, 449-459.
Martin, R. A. (2005). Conservation of freshwater and euryhaline elasmobranchs: a review. Journal of the Marine Biological Association of the United Kingdom 85, 1049-1073.
Mate, B. R., Rossbach, K. A., Nieukirk, S. L., Wells, R. S., Irvine, A. B., Scott, M. D. and Read, A. J. (2006). Satellite-monitored movements and dive behavior of a bottlenose dolphin (Tursiops truncates) in Tampa Bay, Florida. Marine Mammal Science 11, 452-463.
Matern, S. A., Cech, J. J. and Hopkins, T. E. (2000). Diel movements of bat rays, Myliobatus californica, in Tomales Bay, California: evidence for behavioural thermogregualtion? Environmental Biology of Fishes 58, 173-182.
McAuley, R., Lenanton, R., Chidlow, J., Allison, R. and Heist, E. (2005). Biology and stock assessment of the thickskin (sandbar) shark, Carcharhinus plumbeus, in Western Australia and further refinement of the dusky shark, Carcharhinus obscures, stock assessment. Final FRDC Report – Project 2000/13, Fisheries Research Report No. 151, Department of Fisheries, Western Australia.
McAuley, R. B., Simpfendorfer, C. A. and Wright, I. W. (2007). Gillnet mesh selectivity of the sandbar shark (Carcharhinus plumbeus): implications for fisheries management. ICES Journal of Marine Science 64, 1702-1709.
182 McElroy, W. D., Wetherbee, B. M., Mostello, C. S., Lowe, C. G., Crow, G. L. and Wass, R. C. (2006). Food habits and ontogenetic changes in the diet of the sandbar shark Carcharhinus plumbeus, in Hawaii. Environmental Biology of Fishes 76, 81-92.
McGlennon, D. Jones, G. K., Baker, J., Jackson, W. B. and Kinloch, M. A. (2000). Ageing, catch-at-age and relative year-class strength for snapper (Pagrus auratus) in northern Spencer Gulf, South Australia. Marine and Freshwater Research 51, 669-677.
McKibben, J. N. and Nelson, D. R. (1986). Patterns of movement and grouping of gray reef sharks, Carcharhinus amblyrhynchos, at Enewetak, Marshall Islands. Bulletin of Marine Science 38, 89-110.
Meade, R. H. (1966). Salinity variations in the Connecticut River. Water Resources Research 2, 567-579.
Medved, R. J. and Marshall, J. A. (1983). Short-term movements of young sandbar sharks, Carcharhinus plumbeus (Pisces, Carcharhinidae). Bulletin of Marine Science 33, 87-93.
Merson, R. R. and Pratt, H. L. Jr. (2001). Distribution, movements and growth of young sandbar sharks, Carcharhinus plumbeus, in the nursery grounds of Delaware Bay. Environmental Biology of Fishes 61, 31-24.
Mills, C. A. and Mann, R. H. K. (1985). Environmental-induced fluctuations in year- class strength and their implications for management. Journal of Fish Biology 27, 209-226.
Miner, J. G. and Stein, R. A. (1996). Detection of predators and habitat choice by small bluegills: effects of turbidity and alternative prey. Transactions of the American Fisheries Society 125, 97-103.
Morgan, D., Allen, M., Bedford, P. and Horstman, M. (2002). Inland fish fauna of the Fitzroy River Wesetern Australia (including the Bunuba, Gooniyandi, Ngarinyin, Nyikinaand Walmajarri names). Report to the Natural Heritage Trust, Australian Government, Canberra.
Morgan, D. L, Allen, M. G., Bedford, P. and Horstman, M. (2004). Fish fauna of the Fitzroy River in the Kimberley region of Western Australia – including the Bunuba, Gooniyandi, Ngarinyin, Nyikina and Walmajarri Aboriginal names. Records of the Western Australian Museum 22, 147-161.
Morgan, D. L., Ebner, B. C., Beatty, S. J. and Whitty, J. M. (2008). Fish fauna of Snake (Uralla) Creek, Fitzroy River, Kimberley, Western Australia. Report to Liveringa Pastoral Company, Derby, Western Australia.
183 Morgan, D., Thorburn, D., Fenton, J., Wallace-Smith, H. and Goodson, S. (2005). Influence of the Camballin Barrage on fish communities in the Fitzroy River, Western Australia. Report to Land and Water Australia, Australian Government, Canberra.
Morgan, D. L., Whitty, J. M., Phillips, N. M., Thorburn, D. C., Chaplin, J. A. and McAuley, R. (2011). North-western Australia as a hotspot for endangered elasmobranchs with particular reference to sawfishes and the northern river shark. Journal of the Royal Society of Western Australia 94, 345-358.
Morrissey, J. F. and Gruber, S. H. (1993). Habitat selection by juvenile lemon sharks, Negaprion brevirostris. Environmental Biology of Fishes 38, 311-319.
Murphy, D. D. and Noon, B. R. (1992). Integrating scientific methods and habitat conservation planning: reserve design for northern spotted owls. Ecological Applications 2, 4-17.
Natanson, L. J. (1993). Effect of temperature on band deposition in the little Skate, Raja erinacea. Copeia 1993, 199-206.
Nelson, D. R. and Johnson, R. H. (1970). Diel activity rhythms in the nocturnal, bottom-dwelling Heterodontus francisci and Cephaloscyllium ventriosum. Copeia 1970, 732-739.
Ng, C. L., Able, K. W. and Grothues, T. M. (2007). Habitat use, site fidelity, and movement of adult striped bass in a southern New Jersey estuary based on mobile acoustic telemetry. Transactions of the American Fisheries Society 136, 1344-1355.
North, J. A., Beamesderfer, R. C. and Rien, T. A. (1993). Distribution and movements of white sturgeon in the Columbia River Reservoirs. Northwest Science 67, 105-111.
Oliveira, K. (1999). Life history characteristics and strategies of the American eel, Anguilla rostrata. Canadian Journal of Fisheries and Aquatic Sciences 56, 795-802.
Ortega, L. A., Heupel, M. R., Van Benyen, P. and Motta, P. J. (2009). Movement patterns and water quality preferences of juvenile bull sharks (Carcharhinus leucas) in a Florida estuary. Environmental Biology of Fishes 84, 361-373.
Papastamatiou, Y. P., Wetherbee, B. M., Lowe, C. G. and Crow, G. L. (2006). Distribution and diet of four species of carcharhinid shark in the Hawaiian Islands: evidence for resource partitioning and competitive exclusion. Marine Ecology Progress Series 320, 239-251.
Pauly, D. (1980). On the interrelationships between natural mortality, growth parameters and mean environmental temperature in 175 fish stocks. Journal du Conseil - Conseil International pour l’Exploration de la Mer 39, 175–192.
184 Pauly, D. (1987). On the interrelationships between natural mortality, growth parameters, and mean environmental temperature in 175 fish stocks. ICES Journal of Marine Science 39, 175-192.
Pelc, R. and Fujita, R. M. (2002). Renewable energy from the ocean. Marine Policy 26, 471-479.
Peters, H. (1997). Observations of stratified turbulent mixing in an estuary: neap-to- spring variations during high river flow. Estuarine, Coastal and Shelf Science 45, 69-88.
Peterson, I. and Wroblewski, J. S. (1984). Mortality rates of fishes in the pelagic ecosystem. Canadian Journal of Fisheries and Aquatic Sciences 41, 1117- 1120.
Peverell, S. C. (2005). Distribution of sawfishes (Pristidae) in the Queensland Gulf of Carpentaria, Australia, with notes on sawfish ecology. Environmental Biology of Fishes 73: 391-402.
Peverell, S. C. (2008). Sawfish (Pristidae) of the Gulf of Carpentaria, Queensland Australia. MSc Thesis, James Cook University.
Phillips, N. M., Chaplin, J. A., Morgan, D. L. and Peverell, S. C. (2011). Population genetic structure and genetic diversity of three critically endangered Pristis sawfishes in Australian waters. Marine Biology 158, 903-915.
Pikitch, E. K., Chapman, D. D., Babcock, E. A. and Shivji, M. S. (2005). Habitat use and demographic population structure of elasmobranchs at a Caribbean atoll (Glover’s Reef, Belize). Marine Ecology Progress Series 302, 187-197.
Pilgrim, J. M., Fang, X. and Stefan, H. G. (1998). Stream temperature correlations with air temperatures in Minnesota: implications for climate warming. Journal of the American Water Resources Association 34, 1109-1121.
Pillans, R. D., Stevens, J. D., Kyne, P. M. and Salini, J. (2009). Observations on the distribution, biology, short-term movements and habitat requirements of river sharks Glyphis spp. in northern Australia. Endangered Species Research 10, 321-332.
Pogonoski, J. J. and Pollard, D. A. (2003). ‘Glyphis garricki. IUCN Red List of Threatened Species: Glyphis garricki (Version 2009.2)’. Available at www.iucnredlist.org [accessed 4 December 2009].
Pogonoski, J. J., Pollard, D. A. and Paxton, J. R. (2002). Conservation overview and action plan for Australian threatened and potentially threatened marine and estuarine fishes. Environment Australia, Canberra.
Pope, K. L. and Willis, D. W. (1996). Seasonal influences on freshwater fisheries sampling data. Reviews in Fisheries Science 4, 57-73.
185 Power, M. E., Matthews, W. J. and Stewart, A. J. (1985). Grazing minnows, piscivorous bass, and stream algae: dynamics of a strong interaction. Ecology 66, 1448-1456.
Prathumratana, L., Sthiannopkao, S. and Kim, K. W. (2008). The relationship of climatic and hydrological parameters to surface water quality in the lower Mekong River. Environment International 34, 860-866.
Pratt, H. L. Jr. and Carrier, J. C. (2001). A review of elasmobranch reproductive behaviour with a case study on the nurse shark, Ginglymostoma Cirratum. Environmental Biology of Fishes 60, 157-188.
Priede, I. G. (1984). A basking shark (Cetorhinus maximus) tracked by satellite together with simultaneous remote sensing. Fisheries Research 2, 201-216.
Queiroz, N., Lima, F. P., Maia, A., Ribeiro, P. A., Correia, J. P. and Santos, A. M. (2005). Movement of blue shark, Prionace glauca, in the north-east Atlantic based on mark-recapture data. Journal of the Marine Biological Association of the United Kingdom 85, 1107-1112.
Raymond, H. L. (1979). Effects of dams and impoundments on migrations of juvenile chinook salmon and steelhead from the Snake River, 1966 to 1975. Transactions of the American Fisheries Society 108, 505-529.
Renaud, M. L. and Carpenter, J. A. (1994). Movements and submergence patterns of loggerhead turtles (Caretta caretta) in the Gulf of Mexico determined through satellite telemetry. Bulletin of Marine Science 55, 1-15.
Robertson, A. I. and Duke, N. C. (1987). Mangroves as nursery sites: comparisons of the abundance and species composition of fish and crustaceans in mangroves and other nearshore habitats in tropical Australia. Marine Biology 96, 193- 205.
Rosenberg, A. A., Beddington, J. R. and Bason, M. (1986). The growth and longevity of krill during the first decade of pelagic whaling. Nature 324, 152-154.
Sanford, L. P., Sellner, K. G. and Breitburg, D. L. (1990). Covariability of dissolved oxygen with physical processes in the summertime Chesapeake Bay. Journal of Marine Research 48, 567-590.
Schaefer, J. (2001). Riffles as barriers to interpool movement by three cyprinids (Notropis boops, Campostoma anomalum and Cyprinella venusta). Freshwater Biology 46, 379-388.
Scheuerell, M. D. and Schindler, D. E. (2003). Diel vertical migration by juvenile sockeye salmon: empirical evidence for the antipredation window. Ecology 84, 1713-1720.
Schlosser, I. J. (1987). The role of predation in age- and size-related habitat use by stream fishes. Ecology 68, 651-659.
186 Schoener, T. W. (1974). Resource partitioning in ecological communities. Science 185, 27-39.
Schoener, T. W. (1968). The anolis lizards of Bimini: resource partitioning in a complex fauna. Ecology 49, 704-726.
Semeniuk, V. (1980). Mangrove zonation along an eroding coastline in King Sound, north-western Australia. Journal of Ecology 68, 789-812.
Semeniuk, V. (1981). Long-term erosion of the tidal flats King Sound, north Western Australia. Marine Geology 43, 21-48.
Semeniuk, V. (1983). Mangrove distribution in Northwestern Australia in relationship to regional and local freshwater seepage. Plant Ecology 53, 11-31.
Simpfendorfer, C. A. (2000). Predicting population recovery rates for endangered western Atlantic sawfishes using demographic analysis. Environmental Biology of Fishes 58, 371-377.
Simpfendorfer, C. A., Freitas, G. G., Wiley, T. R. and Heupel, M. R. (2005). Distribution and habitat partitioning of immature bull sharks (Carcharhinus leucas) in a southwest Florida estuary. Estuaries 28, 78-85.
Simpfendorfer, C. A., Heupel, M. R. and Collins, A. B. (2008a). Variation in the performance of acoustic receivers and its implication for positioning algorithms in a riverine setting. Canadian Journal of Fisheries and Aquatic Sciences 65, 482-492.
Simpfendorfer, C. A., Poulakis, G. R., O’Donnell, P. M. and Wiley, T. R. (2008b). Growth rates of juvenile smalltooth sawfish Pristis pectinata Latham in the western Atlantic. Journal of Fish Biology 72, 711-723.
Simpfendorfer, C. A. and Wiley, T. R. (2005). Identification of priority areas for smalltooth sawfish conservation. Report to the National Fish and Wildlife Foundation, Washington D.C.
Simpfendorfer, C. A., Wiley, T. and Yeiser, B. G. (2010). Improving conservation planning for an endangered sawfish using data from acoustic telemetry. Biological Conservation 143, 1460-1469.
Simpfendorfer, C. A., Yeiser, B. G., Wiley, T. R., Poulakis, G. R., Stevens, P. W. and Heupel, M. R. (2011). Environmental influences on the spatial ecology of juvenile smalltooth sawfish (Pristis pectinata): results from acoustic monitoring. PLoS ONE 6, e16918.
Simpson, J. H., Brown, J., Matthews, J. and Allen, G. (1990). Tidal straining, density currents, and stirring in the control of estuarine stratification. Estuaries and Coasts 13, 125-132.
187 Sims, D. W. (2003). Tractable models for testing theories about natural strategies: foraging behaviour and habitat selection of free-ranging sharks. Journal of Fish Biology 63, 53-73.
Sims, D. W., Nash, J. and Morritt, D. (2001). Movements and activity of male and female dogfish in a tidal sea lough: alternative behavioural strategies and apparent sexual segregation. Marine Biology 139, 1165-1175.
Sims, D. W., Southall, E. J., Richardson, A. J., Reid, P. C. and Metcalfe, J. D. (2003). Seasonal movements and behaviour of basking sharks from archival tagging: no evidence of winter hibernation. Marine Ecology Progress Series 248, 187- 196.
Sims, D. W., Weathermouth, V. J., Southall, E. J., Hill, J. M., Moore, P., Rawlinson, K., Hutchinson, N., Budd, G. C., Righton, D., Metcalfe, J. D., Nash, J. P. and Morritt, D. (2006). Hunt warm, rest cool: bioenergetic strategy underlying diel vertical migration of a benthic shark. Journal of Animal Ecology 75, 179-190.
Sinokrot, B. A. and Gulliver, J. S. (2000). In-stream flow impact on river water temperatures. Journal of Hydraulic Research 38, 339-350.
Skomal, G. B., Zeeman, S. I., Chisholm, J. H., Summers, E. L., Walsh, H. J., McMahon, K. W. and Thorrold, S. R. (2009). Transequatorial migrations by basking sharks in the Western Atlantic Ocean. Current Biology 19, 1019- 1022.
Smith, A. K. and McPherson, B. P. (1981). Comparison of T-bar and Petersen disc tag loss from chinook salmon (Oncorynchus tshawytscha) in the Rogue River, Oregon. Report to the U.S. Corps of Engineers, Oregon Department of Fish and Wildlife, Portland.
Smith, J. W. and Merriner, J. V. (1985). Food habits and feeding behaviour of the cownose ray, Rhinoptera bonasus, in lower Chesapeake Bay. Estuaries and Coasts 8, 305-310.
Smith, J. W. and Merriner, J. V. (1987). Age and growth, movements and distribution of the cownose ray, Rhioptera bonasus, in Chesapeake Bay. Estuaries 10, 153-164.
Smith, K. and Lavis, M. E. (1975). Environmental influences on the temperature of a small upland stream. OIKOS 26, 228-236.
Southall, E. J., Sims, D. W., Metcalfe, J. D., Doyle, J. I., Franshawe, S., Lacey, C., Shrimpton, J., Solandt, J. L. and Speedie, C. D. (2005). Spatial distribution patterns of basking sharks on the European shelf: preliminary comparison of satellite-tag geolocation, survey and public sightings data. Journal of the Marine Biological Association of the UK 85, 1083-1088.
188 Stanford, J. A., Ward, J. V., Liss, W. J., Frissell, C. A., Williams, R. N., Lichatowich, J. A. and Coutant, C. C. (1996). A general protocol for restoration of regulated rivers. Regulated Rivers: Research and Management 12, 391-413.
Stanley D. W. and Nixon, S. W. (1992). Stratification and bottom-water hypoxia in the Pamlico River estuary. Estuaries and Coasts 15, 270-281.
Stevens, J. D., Bradford, R. W. and West, G. J. (2010). Satellite tagging of blue sharks (Prionace glauca) and other pelagic sharks off eastern Australia: depth behaviour, temperature experience and movements. Marine Biology 157, 575- 591.
Stevens, J. D., McAuley, R. B., Simpfendorfer, C. A. and Pillans, R. D. (2008). Spatial distribution and habitat utilization of sawfish (Pristis spp) in relation to fishing in northern Australia. Report to Department of the Environment, Water, Heritage and the Arts, Australian Government, Canberra.
Stevens, J. D. and McLoughlin K. J. (1991). Distribution, size and sex composition, reproductive biology and diet of sharks from Northern Australia. Australian Journal of Marine and Freshwater Research 42, 151-199.
Stevens, J. D., Pillans, R. D. and Salini, J. (2005). Conservation assessment of Glyphis sp. A (speartooth shark), Glyphis sp. C (northern river shark), Pristis microdon (freshwater sawfish) and Pristis zijsron (green sawfish). Final report to the Department of Environment and Heritage, Australian Government, Canberra.
Stevens, J. D., West, G. J. and McLoughlin, K. J. (2000). Movements, recapture patterns, and factors affecting the return rate of carcharhinid and other sharks tagged off northern Australia. Marine and Freshwater Research 51, 127-141.
Storey, R. H., Froend, P. M. and Froend, R. H. (2001). Fitzroy River system: environmental values. Report prepared for the Waters and Rivers Commission, Perth.
Stuart, I. G. and Berghuis, A. P. (2002). Upstream passage of fish through a vertical- slot fishway in an Australian subtropical river. Fisheries Management and Ecology 9, 111-122.
Student (1908). The probable error of the mean. Biometrika 6, 1-25.
Sundström, L. F., Gruber, S. H., Clermont, S. M., Correia, J. P. S., de Marignac, J. R. C, Morrissey, J. F., Lowrance, C. R., Thomassen, L. and Oliveira, M. T. (2001). Review of elasmobranch behavioral studies using ultrasonic telemetry with special reference to the lemon shark, Negaprion brevirostris, around Bimini Island, Bahamas. Environmental Biology of Fishes 60, 225-250.
Tanaka, S. (1991). Age estimation of freshwater sawfish and sharks in northern Australia and Papua New Guinea. The University Museum, University of Tokyo, Tokyo, Nature and Culture 3, 71-82.
189 Taniuchi, T., Sano, M., Baba, O. and Last, P. R. (1991). Descriptions of freshwater elasmobranchs collected from three rivers in Northern Australia. The University Museum University of Tokyo: Tokyo, Nature and Culture 3, 11-26.
Tavemy, C., Lepage, M., Piefort, S., Dumont, P. and Rochard, E. (2002). Habitat selection by juvenile European sturgeon Acipenser sturio in the Gironde estuary (France). Journal of Applied Ichthyology 18, 536-541.
Taylor, C. F. H. (1999). The role of overbank flow in governing the form of an anabranching river: the Fitzroy River, northwestern Australia. In ‘Fluvial Sedimentology VI’. (Eds N. D. Smith and J. Rogers). pp. 77-91. (Blackwell Science Ltd: Oxford, UK).
Taylor, R. G., Whittington, J. A., Pine, W. E. III and Pollock, K. H. (2006). Effect of different reward levels on tag reporting rates and behaviour of common snook anlgers in southeast Florida. North American Journal of Fisheries Management 26, 645-651.
Teo, S. L. H., Boustany, A., Blackwell, S., Walli, A., Weng, K. C. and Block, B. A. (2004). Validation of geolocation estimates based on light level and sea surface temperature from electronic tags. Marine Ecology Progress Series 283, 81-98.
Theberge, M. M. and Dearden, P. (2006). Detecting a decline in whale shark Rhincodon typus sightings in the Andaman Sea, Thailand, using ecotourist operator-collected data. Oryx 40, 337-342.
Thomas, M. C. and Sheldon, F. (2000). Water resource development and hydrological change in a large dryland river: the Barwon-Darling River, Australia. Journal of Hydrology 228, 10-21.
Thorburn, D. C. (2006). Biology, ecology and trophic interactions of elasmobranchs and other fishes in riverine waters of northern Australia. PhD Thesis, Murdoch University.
Thorburn, D. C. and Morgan D. L. (2004). The northern river shark Glyphis sp. C (Carcharhinidae) discovered in Western Australia. Zootaxa 685, 1-8.
Thorburn, D. C. and Morgan, D. L. (2006). Preliminary survey of freshwater sawfish and other fishes of the McArthur River, Northern Territory. Report to Xstrata, Albion.
Thorburn, D. C., Morgan, D. L., Gill, H., Johnson, M., Wallace-Smith, H., Vigilante, T., Croft, I. and Fenton, J. (2004). Biology and cultural significance of the freshwater sawfish (Pristis microdon) in the Fitzroy River, Kimberley, Western Australia. Report to the Threatened Species Network.
Thorburn, D. C., Morgan, D. L., Rowland, A. J. and Gill, H. S. (2007). Freshwater sawfish Pristis microdon Latham, 1794 (Chondrichthyes: Pristidae) in the Kimberley region of Western Australia. Zootaxa 1471, 27-41.
190 Thorburn, D. C., Morgan, D. L., Rowland, A. J., Gill, H. S. and Paling, E. (2008). Life history notes of the critically endangered dwarf sawfish, Pristis clavata, Garman 1906, from the Kimberley region of Western Australia. Environmental Biology of Fishes 83, 139-145.
Thorburn, D. C., Peverell, S., Stevens, J. D., Last, P. R. and Rowland, A. J. (2003). Status of freshwater and estuarine elasmobranchs in northern Australia. Final Report to Natural Heritage Trust, Australian Government, Canberra.
Thorburn, D. C. and Rowland, A. J. (2008). Juvenile bull sharks Carcharhinus leucas (Valenciennes, 1839) in northern Australian rivers. The Beagle, Records of the Museums and Art Galleries of the Northern Territory 24, 79-86.
Thorson, T. B. (1971). Movement of bull sharks, Carcharhinus leucas, between Caribbean Sea and Lake Nicaragua demonstrated by tagging. Copei 1971, 336-338.
Thorson, T. B. (1982). Life history implications of a tagging study of the largetooth sawfish, Pristis perotteti, in the Lake Nicaragua-Rio San Juan system. Environmental Biology of Fishes 7, 207-228.
Toro, D. M. D. (1978). Optics of turbid estuarine waters: approximations and applications. Water Research 12, 1059-1068.
Torres, L. G., Heithaus, M. R. and Bryan, D. (2006). Influence of teleost abundance on the distribution and abundance of sharks in Florida Bay, USA. Hydrobiologia 569, 449-455.
UNEP-WCMC (2007). ‘CITES species database: Pristis microdon’. Available at http://www.cites.org/eng/resources/species.html [accessed 12 April 2010].
Vincent, C., McConnell, B. J., Ridoux, V. and Fedak, M. A. (2002). Assessment of Argos location accuracy from satellite tags deployed on captive grey seals. Marine Mammal Science 18, 156-166.
Voegeli, F. A., Lacroix, G. L. and Anderson, J. M. (1998). Development of miniature pingers for tracking Atlantic salmon smolts at sea. Hydrobiologia 371/372, 35- 46.
Voegeli, F. A., Smale, M. J., Webber, D. M., Andrade, Y. and O’Dor, R. K. (2001). Ultrasonic telemetry, tracking and automated monitoring technology for sharks. Environmental Biology of Fishes 60, 267-281.
Watsky, M. A. and Gruber, S. H. (1990). Introduction and duration of tonic immobility in the lemon shark Negaprion brevirostris. Fish Physiology and Biochemistry 8, 207-210.
Webb, B. W. and Nobilis, F. (2007). Long-term changes in river temperature and the influence of climatic and hydrological factors. Hydrological Sciences Journal 52, 74-58.
191 Webb, B. W. and Zhang, Y. (1997). Spatial and seasonal variability in the components of the river heat budget. Hydrological processes 11, 79-101.
Weng, K. C., Boustany, A. M., Pyle, P., Anderson, S. D., Brown, A. and Block, B. A. (2007). Migration and habitat of white sharks (Carcharodon Carcharias) in the eastern Pacific Ocean. Marine Biology 152, 877-894.
Weng, K. C., Castilho, P. C., Morrissette, J. M., Landeira-Fernandex, A. M., Holts, D. B., Schallert, R. J., Goldman, K. J. and Block, B. A. (2005). Satellite tagging and cardiac physiology reveal niche expansion in salmon sharks. Science 310, 104-106.
Weng, K., Foley, D. G., Ganong, J. E., Perle, C., Shilinger, G. L. and Block, B. A. (2008). Migration of an upper trophic level predator, the salmon shark (Lamna ditropis), between distant ecoregions. Marine Ecology Progress Series 372, 253-264.
Werner, E. E., Mittelback, G. G., Hall, D. J. and Gilliam, J. F. (1983). Experimental tests of optimal habitat use in fish: the role of relative habitat profitability. Ecology 64, 1525-1539.
West, G. J. and Stevens, J. D. (2001). Archival tagging of school shark, Galeorhinus Galeus, in Australia: initial results. Environmental Biology of Fishes 60, 283- 298.
White, W. T. and Sommerville, E. (2010). Elasmobranchs of Tropical Marine Ecosystems. In ‘Sharks and their relatives II: biodiversity, adaptive physiology, and conservation’. (Eds J. C. Carrier, J. A. Musick and M. R. Heithaus). pp. 159-200. (Taylor and Francis Group: Boca Raton, Florida).
Wirsing, A. J., Heithaus, M. R. and Dill, L. M. (2007). Fear factor: do dugongs (Dugong dugon) trade food for safety from tiger sharks (Galeocerdo cuvier)? Oecologia 153, 1031-1040.
Wolanski, E., Moore, K., Spagnol, S., D’Adamo, N. and Pattiaratchi, C. (2001). Rapid, human-induced siltation of the macro-tidal Ord River estuary, Western Australia. Estuarine, Coastal and Shelf Science 53, 717-732.
Wolanski, E. and Spagnol, S. (2003). Dynamics of the turbidity maximum in King Sound, tropical Western Australia. Estuarine, Coastal and Shelf Science 56, 877-890.
Wong, G. T. F. (1979). Alkalinity and pH in the southern Chesapeake Bay and the James River estuary. Limnology and Oceanography 24, 970-977.
Wueringer, B. E., Squire, L. Jr. and Collin, S. P. (2009). The biology of extinct and extant sawfish (Batoidea: Sclerorhynchidae and Pristidae). Reviews in Fish Biology and Fisheries 19, 445-464.
192 Wyrwoll, K.-H., Hopwood, J. and McKenzie, N. L. (1992). The Holocene paleohydrology and climatic history of the northern great sandy desert-fitzroy trough: with special reference to the history of the northwest Australian monsoon. Climatic Change 22, 47-65.
Xiao, Y., Brown, L. P., Walker, T. I. and Punt, A. (1999). Estimation of instantaneous rates of tag shedding for school shark, Galeorhinus galeus, and gummy shark, Mustelus antarticus, by conditional likelihood. Fishery Bulletin 97, 170-84.
Young, R, Smart, G. and Harding, J. (2004). Impacts of hydro-dams, irrigation schemes and river control works. In ‘Freshwaters of New Zealand’. (Eds J. S. Harding, M. P. Mosely, C. Pearson and B. Sorrell). pp. 37.1-37.15. (New Zealand Hydrological Society and New Zealand Limnological Society: Wellington).
193 Appendix I
Pristis microdon catch data (2005 to 2009)
Tag #, Rototag, acoustic or satellite tag number
TL Tag # Capture date Location Sex (mm) F7056 R 19-May-05 Lower Barrage Pool F > 2660 F7079 13-Jul-05 Lower Barrage Pool F 2299 F7076 14-Jul-05 Lower Barrage Pool M 2059 F7081 3-Aug-06 Lower Barrage Pool F 1150 F7082 3-Aug-06 Lower Barrage Pool F 1050 F7085 3-Aug-06 Lower Barrage Pool M 1070 F7084 6-Aug-06 Camballin Pool F 2230 F7075 6-Aug-06 Myroodah Crossing M 1991 F7068 6-Aug-06 Myroodah Crossing M 1079 F7067 7-Aug-06 Myroodah Crossing F 1012 F7074 9-Aug-06 Snag Pool F 2660 F7072 9-Aug-06 Snag Pool F 2480 F7073 9-Aug-06 Snag Pool F 1320 F7021 R 10-Aug-06 Snag Pool F 2540 F9801 24-Oct-06 Myroodah Crossing M 2100 F9802 25-Oct-06 Lower Barrage Pool M 1051 F9803 26-Oct-06 Snag Pool F 2487 F9804 26-Oct-06 Snag Pool F 2750 F9086 22-Jun-07 Snag Pool F 960 F9836 22-Jun-07 Snag Pool F 1000 F9835 22-Jun-07 Snag Pool F 1555 F9838 22-Jun-07 Snag Pool M 985 F7094 23-Jun-07 Snag Pool F 983 F9840 23-Jun-07 Snag Pool F 1060 F9848 23-Jun-07 Snag Pool F 1024 F9841 23-Jun-07 Snag Pool F 846 F8086 23-Jun-07 Snag Pool M 950 F9810 23-Jun-07 Snag Pool M 822 F9837 23-Jun-07 Snag Pool M 1050 F7092 23-Jun-07 Snag Pool M 1611 F9839 23-Jun-07 Snag Pool M 950 F9807 23-Jun-07 Snag Pool M 941 F9808 23-Jun-07 Snag Pool M 989 F9809 23-Jun-07 Snag Pool M 825 F9847 23-Jun-07 Snag Pool M 911 F9849 23-Jun-07 Snag Pool M 1016 F9850 23-Jun-07 Snag Pool M 1000 F9826 24-Jun-07 Snag Pool F 933 F9842 24-Jun-07 Snag Pool M 941 F7096 29-Jun-07 Camballin Pool M 935 F9843 30-Jun-07 Camballin Pool F 900 F9830 3-Jul-07 Snag Pool F 1780 F9831 3-Jul-07 Snag Pool F 885 F9832 3-Jul-07 Snag Pool F 789 F7093 3-Jul-07 Snag Pool M 936
194 TL Tag # Capture date Location Sex (mm) F9833 3-Jul-07 Snag Pool M 1070 F9834 4-Jul-07 Snag Pool F 886 F9844 4-Jul-07 Snag Pool F 914 M1007 18-Jul-07 Snag Pool F 842 M1013 18-Jul-07 Snag Pool M 1580 M1003 19-Jul-07 Snag Pool F 981 M1006 19-Jul-07 Snag Pool F 990 M1005 19-Jul-07 Snag Pool M 854 SPOTPM1 24-Oct-07 Pt Torment, King Sound M 2580 M1004 1-Nov-07 Camballin Pool F 1111 1038502 13-Nov-07 Langi Crossing F 1576 M1002 24-Jun-08 Myroodah Crossing M 1897 M1044 25-Jun-08 Myroodah Crossing M 879 M1031 27-Jun-08 Camballin Pool M 812 M1022 29-Jun-08 Snag Pool F 918 M1021 29-Jun-08 Snag Pool F 1423 M1038 29-Jun-08 Snag Pool F 873 1038504 29-Jun-08 Snag Pool M 1429 M1023 30-Jun-08 Snag Pool F 959 M1032 30-Jun-08 Snag Pool F 977 M1027 30-Jun-08 Snag Pool M 852 M1043 1-Jul-08 Snag Pool F 880 M1035 1-Jul-08 Snag Pool F 989 NA 23-Jul-08 Snag Pool M 1934 M1048 27-Oct-08 Myroodah Crossing F 2142 M1045 28-Oct-08 Camballin Pool M 2210 M1049 31-Oct-08 Snag Pool F 2090 M1047 31-Oct-08 Snag Pool M 2102 M1051 1-Nov-08 Snag Pool M 1883 DOA 18-Nov-08 Telegraph Pool F 925 M1086 12-May-09 Snag Pool F 837 M1077 12-May-09 Snag Pool M 965 M1083 12-May-09 Snag Pool M 763 M1088 12-May-09 Snag Pool M 934 NA 12-May-09 Snag Pool F 796 M1076 13-May-09 Camballin Pool F 952 M1090 13-May-09 Camballin Pool F 923 M1078 13-May-09 Camballin Pool F 1027 M1082 13-May-09 Camballin Pool F 867 NA 13-May-09 Camballin Pool F 940 M1084 14-May-09 Camballin Pool M 1016 M1066 14-May-09 Lower Barrage Pool F 1013 M1081 14-May-09 Lower Barrage Pool - 911 M1079 14-May-09 Lower Barrage Pool F 872 M1087 15-May-09 Lower Barrage Pool M 901 M1098 16-Sep-09 Myroodah Crossing M 1003 M1092 16-Sep-09 Myroodah Crossing M 968 M1034 17-Sep-09 Myroodah Crossing M 1865 M1093 15-Sep-09 Myroodah Crossing M 1580 M1094 17-Sep-09 Myroodah Crossing F 2499 M1095 22-Sep-09 Myroodah Crossing M 1065
195
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