Feasibility of reef restoration in a south-western Australian estuary

This thesis is presented for the degree of Bachelor of Science Honours, College of Science, Health, Engineering and Education, of Murdoch University, 2019

Lauren Peck (BSc)

Declaration

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

Lauren Peck

Abstract

With 85% of reefs lost around the world within the last 130 years, these reefs are now one of the most threatened marine habitats in the world and in less than

10% of naturally occurring oyster reefs remain. Shellfish reefs provide a range of services that promote healthy ecosystems, including water filtration, fish production and shoreline erosion. In estuaries, these services are extremely important as human activities are increasing degrading these environments. Thus, shellfish reefs can aid in restoring ecosystem functioning of an estuary while providing additional ecosystem services. The aim of this study was to determine the feasibility of a number of shellfish reef options for the Peel-Harvey Estuary in south-. This first involved exploring the historical and current distributions of shellfish to elucidate whether shellfish reefs existed in the Peel-Harvey Estuary and to identify a suite of candidate species.

A bioclimatic modelling approach was then used to elucidate the suitability of five native

Australian oyster species to the environmental conditions that occur in the Peel-Harvey

Estuary, the largest estuary in south-western Australia. Laboratory tank trials were then used to validate the results of that model, in which the two most suitable species, i.e.

Ostrea angasi and , were exposed, for two months, to the extremes in water temperature and salinity that occur in the Peel-Harvey Estuary during summer (26◦C and 48ppt) and winter (15◦C and 14ppt) and their survival, body condition index (BCI) and behaviour (valve activity) monitored.

The probability of survival (S) over the duration of the study was lowest for O. angasi in the summer (S=0.0) and winter treatments (S=0.18), which both differed significantly

(P<0.001) from the control group (marine condition), and were consistent with the salinity in the winter and summer treatments falling outside previously recorded tolerance thresholds. In contrast, S. glomerata had a high probability of survival in winter (S=0.98), and ~50% survived the extreme summer conditions. A significant difference of valve activity was found for O. angasi between the three conditions (P<0.05), however, only a

iii

significant difference in valve activity for S. glomerata was found between the marine and summer (P<0.05), and marine and winter conditions (P<0.05). Overall, body condition index (BCI) did not differ significantly (P>0.05) before or among treatments.

The results of the bioclimatic model and survival analyses suggest that S. glomerata was the most suitable candidate for shellfish reef restoration in the Peel-Harvey Estuary, given in particular the extremes that occur in salinity. Further work is required to determine the most suitable areas in the estuary that would maximise survival and growth and thus where such reefs would have a positive impact on the overall health and resilience of the Peel-Harvey Estuary.

iv

Contents Abstract ...... iii Acknowledgements ...... vi Chapter 1: General introduction ...... 1 1.1 Importance of oyster reefs ...... 1 1.2 Threats to oyster reefs ...... 2 1.3 Monitoring stress in ...... 6 1.4 ...... 7 1.5 Australian reef-forming oyster species ...... 8 1.5.1 Potential for filter feeders to ameliorate degradation of south-western Australian estuaries ...... 11 1.6 Aims ...... 12 Chapter 2: Materials and Methods ...... 13 2.1. Peel-Harvey Estuary ...... 13 2.2 Historical and current distribution of shellfish in the Peel-Harvey Estuary ...... 17 2.3 Candidate shellfish species selection ...... 19 2.4 Laboratory experiments ...... 19 2.4.1 Feeding ...... 21 2.4.2 Survival, body condition and valve activity analyses ...... 22 Chapter 3: Results ...... 26 3.1 Historical and current distribution of reef forming shellfish in the Peel-Harvey Estuary ...... 26 3.2 Environmental parameters & species selection ...... 29 3.3 Laboratory Trial ...... 32 3.3.1 Survival analyses ...... 32 3.3.2 Valve Activity ...... 36 3.3.3 Body Condition Index ...... 37 Chapter 4: Discussion ...... 39 4.1 Distribution of oysters in the Peel-Harvey Estuary ...... 39 4.2 Survivability and physiological responses under different environmental conditions 41 4.3 The ecological and social risk of oyster reef restoration in the Peel-Harvey Estuary 44 4.4 The potential of restoration in the Peel-Harvey Estuary using oyster reefs ...... 45 Chapter 5: Conclusions ...... 47 5.1 Future recommendations for oyster reef restoration in the Peel-Harvey Estuary...... 47 5.2 Concluding remarks ...... 48 References ...... 49

v

Acknowledgements

Firstly, I would like thank my supervisors, Alan Cottingham and James Tweedley, for their on-going support and guidance throughout this process. This would not have been possible without Alan helping to set up the tank system, and always being there to help when something goes wrong. Thank you both for always providing feedback on my drafts, guidance through analyses, and for developing my writing and research skills.

Thank you to the Blue Lagoon Farm and the Albany Shellfish Hatchery for the generous donation of oysters, which without, this project could not have been possible.

To the Harry Butler Institute, thank you for the scholarship award, this has provided me freedom to be able to solely focus on my thesis and not stress about finances.

Lastly, thank you to my family, my partner, Joe, and close friends, for the continuous support throughout the last 18 months. To my parents, this would not have been possible without your on-going support, love and faith. I wouldn’t be where I am today, without you all.

vi

Chapter 1: General introduction

1.1 Importance of oyster reefs

Oyster reefs are distributed in intertidal and subtidal waters throughout a range of aquatic habitats, from freshwater to coastal environments (Guo et al., 2015). These habitats can occur in a range of physiochemical conditions, from the relatively stable marine environment to the fluctuating estuarine environment (Akberali & Trueman, 1985). In estuaries, water temperature and salinity are major limiting factors that determine shellfish distribution (Heilmayer et al., 2008; Munroe et al., 2013; Lowe et al., 2017).

Oysters are not capable of regulating their body temperature or salinity of their body fluids, as a result, their metabolic activity and salt content is influenced by their surrounding environment. The synergistic effect of temperature and salinity affects virtually every aspect of an oyster’s life, influencing their physical extent and physiological rates, which controls their survival, valve activity, body condition, growth, respiration, feeding, and excretion (Kennedy et al., 1996; Heilmayer et al., 2008;

McFarland & Hare, 2018).

As ecosystem engineers, oysters, which cement to hard substrates and to other oyster shells to form large complex three-dimensional structures, play an important role in estuaries by providing a range of services that promote healthy ecosystems (Coen et al.,

2007; Grabowski & Peterson, 2007). Growing in dense accumulations they provide a habitat for a diversity of species. For example, many recreational and commercially important fish and crustaceans, together with smaller epibenthic species, utilize the reef for recruitment substrates, nursery areas and foraging purposes (Harding & Mann, 2001;

Lenihan et al., 2001; Tolley & Volety, 2005) This, in turn, increases fish production and the overall biodiversity of the ecosystem (Peterson et al., 2003; Ruesink et al., 2005;

Rezek et al., 2017). For example, in Mobile Bay, USA, Scyphers et al. (2011) found that restored oyster reefs supported higher abundances and different communities of fish and mobile invertebrates compared to areas without oyster reefs. These restored oyster reef

1

habitats also enhanced abundances of commercially important species such as blue crabs (Callinectes sapidus) (+297%), red drum (Sciaenops ocellatus) (+108%), spotted seatrout (Cynoscion nebulosus) (+88%) and flounder (Paralichthys spp.) (+79%)

(Scyphers et al., 2011).

In addition to the aforementioned ecological benefits, oyster reefs contribute to the regulation of estuarine habitats through sediment stabilization, shoreline protection and water quality maintenance (Meyer et al., 1997). Oyster reefs serve as natural breakwaters absorbing the force of water with their physical structure, in turn reducing erosion and creating calmer waters that subsequently protect other habitats such as marshes and beds (Piazza et al., 2005).

Through their suspension feeding, oysters contribute to water quality by filtering out organic and inorganic particles that directly or indirectly enter the system, thereby reducing eutrophication, enhancing water clarity and light penetration and trapping contaminants (Cloern, 1982; Newell, 1988). For example, the population of cockles

(Cerastoderma edule) and (Mytilus edulis), in the Oosterchelde Estuary, the

Netherlands, has the potential to filter the entire volume of the estuary in as little as four days (Smaal et al., 1986; Dame et al., 1991). Given the residence time of water in the estuary can fluctuate between one and three months, the filtering rate of bivalves significantly help to filter the system of unwanted phytoplankton abundances and other contaminants (Tweedley et al., 2016b).

1.2 Threats to oyster reefs

Oyster reefs are one of the most threatened marine habitats in the world (Beck et al.,

2011). While, in the last 130 years, 85% of naturally occurring oyster reefs have been lost globally, this value is much greater for the losses of 90-99% that have occurred in

Australia (Fig. 1; Beck 2009). Among the five native oyster species that occur in

Australia, the most severe declines occurred in the Australian Flat Oyster (

2

angasi), with less than 1% of historical reef systems remaining, followed by the Sydney

Rock Oyster (Saccostrea glomerata), with 10% of historical reef systems remaining

(Gillies et al., 2018). The source and magnitude of extensive marine habitat degradation

are primarily caused by humans (Geraldi et al., 2013). The major drivers of the loss of

oyster reefs are a combination of overfishing for food and harvesting for lime production

(Pollack et al., 2012; Gillies et al., 2018). Other contributing factors include native

diseases, e.g. unknown (QX), introduced diseases, e.g. Bonamiosis,

invasive species, spionid worms (mud worms) polyclad flat worms as well as habitat loss

(Nell, 2001; Beck, 2009).

Fig. 1.1 The conditions of oyster reefs globally. Condition ratings based on the percentage of current historical abundance of oyster reefs remaining: <50% lost (good); 50-89 % lost (fair); 90-99% lost (poor); >99% (functionally extinct). Taken from (Beck, 2009).

In Australia, both O. angasi, and S. glomerata occurred in vast quantities, however, their

abundances rapidly declined after the European settlement in 1788. The first O. angasi

fishery was established in in the 1840s, but collapsed in 1895 (Beck,

2009). With a total of 6 million oysters reported to have been harvested from a single

location in the Coffin Bay, SA, the heavy fishing activities left the natural beds of

O. angasi depleted by the time legislations were in place to control dredging (Beck,

2009). By 1945 there were no records of fishing being undertaken and only remnants of

the O. angasi population remained. In NSW and Queensland, the extensive beds of 3

S. glomerata suffered the same fate as O. angasi. Large quantities were harvested to produce lime for construction in Sydney, declining wild reefs between the 1850s and

1870s (Nell, 2001). In Queensland, fisheries expanded, exporting an annual 252,000 individual oysters, from a single fishery, to Sydney and Melbourne. Oyster fisheries along the eastern coast of Australia generally peaked in the 1880s, and were exploited by the

1920s, due to over harvesting and the appearance of the parasitic mud worm and diseases (Ogburn et al., 2007). In Albany, Western Australia, Oyster Harbour was an iconic O. angasi oyster reef population, giving the location its name (Beck, 2009; Gillies et al., 2018). However, during the 1850s unsustainable dredge fishing of the harbour occurred and led to the decline of O. angasi oyster populations by 1880 (Saville-Kent,

1893a). Historically, other systems, such as the Hardy Inlet and Swan River Estuary have been known to support large populations of O. angasi, however, changes due to geomorphology and hydrology to the system and/or extensive harvesting by European colonies for food and lime production resulted in the populations collapse (Saville-Kent,

1893b; Brearley & Hodgkin, 2005).

Queensland unknown (QX) disease is caused by the protozoan, Marteilia sydneyi, that naturally occurs along the east coast of Australia, infecting oyster species such as;

S. glomerata, O. angasi, S. echinata and S. forskale. This protozoan may kill up to 80% of infected oysters, by invading the digestive gland, causing the oysters to die of starvation (Hine, 1996). The QX disease has led to the decline of oyster farms in

Australia, for example, an oyster industry in the Georges River, Sydney, collapsed within seven years of QX being detected (Bower & Kleeman, 2011). Bonamia exitosa is a parasite of oysters, more commonly known as the Bonamiosis disease. Bonamiosis generally affects the genus Ostrea, but have also been known to affect virginica and Saccostrea glomerata. Infected oysters may appear normal, others have a yellow discolouration and/or extensive lesions in the connective tissues of the gills, mantle and digestive glands, resulting in mortality (Bower, 2015). A similar disease has led to the mass mortality of the European Flat Oyster, , which produced an

4

annual yield of >50 million oysters, which led to virtual extinction of native oyster beds throughout much of Europe by the early to mid-1900s (Beck, 2009). Of the four spionids that occur, Polydora websteri, is the most damaging, infesting and killing large populations of S. glomerata (Nell, 2001). These polychaetes dwell in the inside of oyster shells, forming mud blisters that if not sealed over by the shell of a healthy oyster, will cause eventual death. Species of spionids including P. websteris, Boccardia knoxi and

Polydora hoplura have been a problem, affecting native and commercial populations of

S. glomerata and O. angasi in NSW, TAS and South Australia (Nell, 2001; Ogburn et al.,

2007). Flatworms, are a serious problem to global populations of oysters and other bivalves. Flatworm, Imogine mcgrathi, was identified in the early 1900s as a threat to oysters in NSW (Nell, 2001). Consuming at a rate of one oyster per flatworm per month, and when occurring in high densities, flatworms can have drastic effects of oyster populations (Nell, 2001).

Habitat loss has been a contributing factor to the loss of native oyster reefs in many regions around the world. The rapid development of coastal industries and cities in

China, has led to the loss of habitat in the past 30 years, removing up to 70% of the

Hangu oyster reef in Bohai Bay and 90% of the Dajiawa oyster reef (Beck, 2009). Further loss of reefs in China, occurred through the bombing of reefs in order to provide access for commercial ships. Loss of oyster reefs, due to coastal development has also occurred throughout Europe. In South America, where the loss of oyster reefs is yet to be as dire as many other regions in the world, the extensive loss of mangroves is causing the loss of several species that use the mangrove as their primary settlement habitat. In particular the mangrove oyster (Crassostrea rhizophorae) has been listed on Colombia’s Red List of threatened species due to the significant population loss in the coastal lagoon,

Cienaga Grande de Santa Marta (Beck, 2009).

In estuaries, fluctuating temperature and salinity has also played a major role in the mortality of oysters throughout the world (Heilmayer et al., 2008; Verdelhos et al., 2015;

5

Cole et al., 2016; Rybovich et al., 2016; Lowe et al., 2017; Taylor et al., 2017). Mortality of many oyster species are typically greatest when exposed to high temperatures, and especially when associated with high or low salinities (Rybovich et al., 2016; Lowe et al.,

2017). This is evident in Crassostrea gigas, which is cultured in many countries and have been suffering from heavy summer mortalities due to unseasonal freshwater discharge during the summer (Guo et al., 2015). Numerous studies showcase the high mortality of oysters due to long term exposure to low salinity during summer months, typically due to tropical storm events bringing an influx of freshwater into the system (Dugas & Roussel,

1983; La Peyre et al., 2013; Munroe et al., 2013; Rybovich et al., 2016). When environments are not favourable, oysters will shut their valves in order to reduce shock to their tissues, however, with a high metabolic rate induced by summer temperatures, oysters need to open in order to feed and respire resulting in the exposure to harsh salinities that cause osmotic shock and eventual death (Akberali & Trueman, 1985).

1.3 Monitoring stress in oysters

In order to survive stressful environments, oysters have adapted the ability to close their valves, temporarily isolating their tissues from the outside environment (Akberali &

Trueman, 1985; Kennedy et al., 1996). A study conducted by Casas et al. (2018a) found that lowered valve openness was detected when the C. virginica were exposed to temperatures of ≤ 10°C and ≥ 30°C, which may be indicative of stressful environments.

When exposed to different salinities, C. virginica typically spent 50-60% of the time with open valves between 6-25ppt salinity and ~30% of the time open at a 3ppt (Casas et al.,

2018b). Closed valves prevent drastic changes in the osmotic concentration of bodily fluids when exposed to short-term fluctuations in salinities. The ability to close valves when exposed to long-term salinity changes may provide a period of time for the oyster to acclimate and prevent osmotic shock (Shumway, 1977). Although valve closure may help the individual survive adverse temporary changes in the environments, it may not improve survival if the changes to the environment are long-term or permanent. With valve closure, the subsequently reduces activities related to respiration,

6

reproduction, feeding and exchanges or gases and metabolites (Akberali & Trueman,

1985).

Body condition index (BCI) compares the meat of the oyster with their theoretical maximum size. A higher value for condition index indicates a greater amount and quality of meat for an oyster. A lowered BCI may serve as a measure of physiological stress to the animal, whether the stress is associated with disease or unfavourable environmental conditions (Mann, 1978). A study conducted by Heilmayer et al. (2008) demonstrated that C. virginica had a low body condition index (BCI) after being exposed to relatively high temperatures (25°C) and low salinities (<5ppt), compared a greater BCI recorded in temperatures and salinities within normal physiological tolerable range (<17.5°C and

15 to 25ppt). Body condition index is a common method to determine the quality and yield of bivalves and fish in . A higher quality of oyster meat tends to weigh more and fill a large volume of the shell, whereas low quality meats are shrunken and contains a higher water content (Haven, 1962; Ricker, 1975; Mason & Nell, 1995).

1.4 Oyster reef restoration

Oyster reef restoration can occur in two ways; firstly, by restoring an oyster reef to its original location, or secondly, to restore oyster reefs in new locations that are suitable for their survival of oysters, in turn restoring the overall habitat of oyster reefs. Restoration of oyster reefs is relatively new, in comparison to that of other marine habitats such as restoration of seagrass meadows and coral reefs. The lack of focus on shellfish reefs could be due to a ‘shifting baseline’, where very few scientists today have witnessed a fully functioning and undisturbed oyster reef and thus the ecosystem services such a habitat can provide (Paling et al., 2001; van Keulen, 2002; Kirby, 2004; Verduin et al.,

2010; Rinkevich, 2014). Concerted efforts in restoring the oyster reefs have been undertaken throughout America, Europe and Australia (Kirby, 2004), showing promise in restoring the natural populations of oyster reefs. Restoration efforts dominantly involve

7

the recycling of used oyster shells to create a new substrate for juvenile oysters to attach to (Naturally Resilient Communities, N.D).

In Australia, oyster restoration is undergoing in Oyster Harbour in Albany, Western

Australia, Windara Reef in the Yorke Peninsula, South Australia and Port Phillip Bay in

Victoria by The Nature Conservancy. In Oyster Harbour, a three-phase plan has been initiated to recover the oyster reefs that were lost in the 1800s during European settlement. Restoring and monitoring an 800 m2 oyster reef, hopes to not only revive the lost ecosystem but increase the water quality and fish stocks that were consequently lost with the oyster reefs. A similar project is currently under way in Yorke Peninsula with a four hectare oyster reef trial, which will be expanded in the coming years to reach 20 hectares, and Port Phillip Bay Victoria, with a 1,200 m2 oyster reef with hopes of expanding the restoration activities to a 500 hectare reef (Jupp, 2019; Nature

Conservancy, 2019a,b).

1.5 Australian reef-forming oyster species

In Australia, there are a number of native reef-forming oyster species, which include the

Australia Flat Oyster, O. angasi, the Sydney , S. glomerata, the Western

Rock Oyster, Saccostrea cucullata, the Coral Oyster, Saccostrea scyphophilla, and the

Tropical Blacklip Oyster, Saccostrea echinata. A brief summary of their distribution and ecology is provided below. Note that there is a paucity of information for S. cucullata,

S. scyphophilla and S. echinata as they are not of commercial importance and have been the subject of less research (Venture, 2016).

Ostrea angasi has historically been recorded in 177 locations along the southern coast of Australia, from northern to the Swan River Estuary in Western

Australia, where it forms reefs or beds on hard or soft substrates in subtidal depths of

< 30 m (Fig. 1.2; Morton & Slack-Smith, 2003; Ogburn et al., 2007). However, today, only one location in harbours a historic O. angasi reef (Gillies et al., 2018). This

8

species is also of major commercial interesting and is currently being cultured in

Tasmania (Menzie et al., 2013), New South Wales (Casas et al., 2018a), Victoria (La

Peyre et al., 2015), South Australia (Saville-Kent, 1893a) and Western Australia (Warton

& Hui, 2011).

Saccostrea glomerata is a (sub)tropical species that is naturally distributed from the

VIC/NSW border and around the northern coast of Australia and as far down the west coast as Shark Bay (Fig. 1.2; Ogburn 2007). This species has previously been recorded at 126 locations on both hard and soft substrates in water depths of >8 m, however, this number has since been reduced to only 6 locations due to disease, overfishing, and dredging. Aquaculture of S. glomerata is currently being undertaken in NSW, QLD and

WA (Gillies et al., 2018). Saccostrea glomerata congener, S. cucullata occurs in estuaries from the and WA as far south as the Swan River Estuary, where it forms beds on hard substrates (Fig. 1.2; Venture, 2016).

Saccostrea scyphophilla and Saccostrea echinata are distributed along the intertidal and shallow subtidal zones of the Northern coasts of Australia (Fig. 1.2). The WA distribution is not well studied for these two species; however, anecdotal evidence suggests that S. scyphophilla shares most of its habitat with S. cucullata but is present in lower densities

(Angell, 1986; Rubio et al., 2013). Saccostrea echinata is prevalent in mangrove estuaries along the northern coast of WA and are harvested recreationally and culturally from the border of the Northern Territory down the west coast to Shark bay, Western

Australia (Hine & Thorne, 2000; Mueller & Woodland, 2015; Venture, 2016).

Due to the Mortality Syndrome (Department of Primary Industries, 2016) affecting populations of Crassostrea gigas in New South Wales, in combination with the species being introduced, this is not currently a suitable species for restoration and thus was not considered for this study (Department of Primary Industries, 2016).

9

Fig. 1.2. The distribution of the chosen five oyster species in Australia.

10

1.5.1 Potential for filter feeders to ameliorate degradation of south-western Australian estuaries

Temperate estuaries are considered the most degraded of all aquatic ecosystems globally (Jackson et al., 2001), with all except one of the 45 estuaries in south-western

Australia being regarded as having been subjected to anthropogenic modification and environmental perturbation; e.g. eutrophication, urbanisation and/or habitat loss

(Jackson et al., 2001; Commonwealth of Australia, 2002; Tweedley et al., 2017).

Moreover, due to their Mediterranean climate, and thus highly seasonal rainfall, and low tidal ranges (i.e. microtidal < 2 m), estuaries in this region are particularly susceptible to the effects of environmental degradation and thus, such anthropogenic impacts that often result in algal blooms and hypoxia events, some of which lead to the mass mortality of marine fauna (Tweedley et al., 2016a, b; Warwick et al., 2018). While, during periods of low rainfall, estuaries in Denmark exhibit similar blooms, however, these systems are fjordic, with hard rocky substrates that support extensive beds of bivalve and ascidians, which significantly reduce phytoplankton concentrations, alleviating harmful events

(Conley et al., 2000). Similarly, annual net phytoplankton production in the Scheldt

Estuary in Netherlands, is negative, reflecting the presence of extensive bivalve beds

(Tweedley et al., 2016b). Such is the filter-feeding capacity of such organisms that in the

Bay of Brest (France) the entire volume of water in the bay is filtered every 96 hours

(Hily, 1991). These international examples demonstrate that increasing the abundance of suspension feeding organisms may help reduce the effects of natural climatic and anthropogenic stressors in south-western Australian estuaries (Warwick et al., 2018;

Wilcox et al., 2018).

Benthic faunal surveys in the sandy and muddy sediments of south-western Australian estuaries have found a lack of bivalves, particularly large suspension-feeders (Wildsmith et al., 2009; 2011; Rose et al., 2019), which could reflect the lack of hard substrates for attachment. Moreover, evidence suggests that the health of some of these systems, in particular their benthic environment has declined over time predominantly due to the

11

deleterious effect of anthropogenic activities (Tweedley et al., 2012; 2014). Given that anthropogenic stressors on these estuaries will only increase with an expanding population and the effects of climate change (Hallett et al., 2018), restoring bivalve populations has been suggested as a potential method of mitigating these effects

(Warwick et al., 2018).

One such estuary in south-western Australia with a long history of anthropogenic perturbation and modification is the Peel-Harvey Estuary (Elliott et al., 2016; Valesini et al., 2019). As it is located next to one of the fastest growing cities in Australia, being highly eutrophic and having relatively low populations of suspension feeding bivalves, this estuary is an ideal candidate to test the theory that the provision of bivalve reefs could improve the health and condition of the system.

1.6 Aims

The objective of this study is to determine whether oyster reef restoration is feasible in the Peel-Harvey Estuary in hopes to contribute to the overall restoration of oyster reef habitats. The specific aims are to:

1. Determine the historical and current distribution of reef-forming shellfish in the

Peel-Harvey Estuary.

2. Determine which reef-forming shellfish species would potentially survive in the

Peel-Harvey Estuary using a bioclimatic modelling approach, i.e. whether the

salinity and temperature tolerance range of each species falls within the

extremes in summer and winter in the Peel-Harvey Estuary.

3. Validate the above modelling approach through measuring survival, valve

activity and body condition on the resultant two best candidate species under

extreme Peel-Harvey ‘summer’ and ‘winter’ conditions replicated in the

laboratory.

12

Chapter 2: Materials and Methods 2.1. Peel-Harvey Estuary The Peel-Harvey Estuary, located 80 km south of the City of in south-western

Australia, is the largest estuary in the region with an area of 131 km2 (McComb, 1995;

Valesini et al., 2019). It comprises two adjoining basins, the Peel Inlet (75 km2) and the

Harvey Estuary (56 km2), which receives freshwater from three rivers; Murray,

Serpentine (Peel Inlet) and Harvey (Harvey Estuary; Fig. 2.1). The estuary has two permanent connections to the Indian Ocean, a 125 m wide and 5 km long shallow natural entrance channel to the north end of the Peel Inlet and the large, deep (200 m wide, 2.5 km long and 4.5-6.5 m depth) artificial Dawesville Channel, located at the northern part of the Harvey Estuary (Department of Marine and Harbours, 1992). The estuary is generally shallow, with average and maximum depths of 0.5 and 2.5 m, respectively.

The catchments of the three rivers have been extensively modified, with 53% of that land cleared for agriculture and a further 2% developed for urban or industrial uses (Valesini et al., 2019). For prosperous agriculture, fertilisers were used to compensate for the low quality of the sandy soils. This led to vast volumes of nutrients being washed into the estuary causing eutrophication. These anthropogenic stresses are compounded by the microtidal tidal range (~0.5 m) and highly seasonal rainfall and freshwater discharge, i.e. with 90 and 95% of the annual quantities occurring between May and October. This resulted in limited flushing and long retention times of water and thus nutrients in the system. In the 1970s and 80s, the estuary became hypereutrophic, manifesting in extensive blooms of macroalgae in the genera Cladophora, Chaetomorpha and Ulva and the toxic cyanobacterium Nodularia spumigena which resulted in mortalities among the less mobile bottom‐living fishes, including crabs, and unpleasant smells (Potter et al.,

1983; McComb, 1995; Steckis et al., 1995).

To combat this, an artificial channel, Dawesville Channel, was opened in 1994, which had significant impacts to the overall health of the estuary (Wildsmith et al., 2009). The

13

most notable changes were seen in the hydrology, water quality and by a shift in ecological composition (Fig. 2.2). The cut significantly increased the tidal range, from 17

– 48% of the ocean tides in the Peel Inlet and 15 – 55% in the Harvey Estuary, and decreased the retention time from 45 days to 22 days in the winter and 68 days to 50 days in the summer (Potter et al., 2016; Valesini et al., 2019). Additionally nutrient concentrations fell rapidly after the opening of the Dawesville Channel, with annual averages of total phosphorus (TP) and total nitrogen (TN), decreasing from 0.15mg/L to

0.05mg/L and 2.0mg/L to 0.5mg/L, respectively, and the high concentrations of chlorophyll-a reduced significantly from >200 ug/L to <10ug/L, reflecting decreasing algal blooms in the basins (Valesini et al., 2019). However, the new channel also increased the overall salinity and the propagation of the salt wedge in the rivers, causing a shift in flora and fauna to more saline tolerant species (Wildsmith et al., 2009; Potter et al.,

2016). Algal blooms and hypoxia have remained evident in the rivers, subsequently leading to mass mortality events of fish. A total of 47 fish kill events occurred during

1999-2017, with kills between 1000 to >10,000 individuals per event. The majority of events occurred in the Murray and Serpentine rivers, due to phytoplankton blooms causing hypoxia (Valesini et al., 2019). The historical battles with anthropogenic stressors to the Peel-Harvey Estuary (Fig. 2.2) thus reinforces the need for restoration of the estuary.

14

Fig. 2.1 The Peel-Harvey Estuary, Western Australia. Map insert shows the location of the estuary in Western Australia.

.

15

Fig. 2.2 Summaries of key environmental, estuary response and socioeconomic changes in the Peel-Harvey system from the early 1800s to present. Taken from (Valesini et al., 2019).

16

2.2 Historical and current distribution of shellfish in the Peel-Harvey Estuary An extensive literature search on the natural history of the Peel region was undertaken to determine if reef-forming bivalves, such as the Australian Flat Oyster Ostrea angasi, occurred in the estuary in the past. This involved a thorough examination of published literature and liaising with the Mandurah Community Museum, Mandurah Licensed

Fishermen’s Association and commercial fishers whose family have fished in the estuary for generations.

Advanced searches in the databases; Google scholar, Aquatic Science and Fisheries

Abstracts, Murdoch University Library and Scopus were undertaken with the following key words: ‘Peel-Harvey Estuary’, ‘South West Australian Estuary’, ‘distribution’,

‘oysters’, ‘benthic faunal survey’, ‘Ostrea angasi’, ‘Saccostrea glomerata’, ‘Saccostrea cucullata’, ‘Saccostrea echinata’ and ‘Saccostrea scyphophilla’.

A visual survey of the Peel-Harvey Estuary was conducted by boat to provide a snap shot of existing oyster species. The Peel Inlet was chosen for visual surveys due to the abundance of man-made hard structures, such as traffic bridges, canals, jetties and pylons and natural hard structures, rocks to which oyster could attach. The surveys were conducted at low tide from a boat and a waterproof camera mounted on an extension pole was used to observe any oysters below the water surface (Fig. 2.3)

17

Fig. 2.3 Oyster survey sites within the Mandurah Channel of the Peel-Harvey Estuary, Perth, Western Australia. Yellow circles and lines highlight the sites that observations were conducted. Taken from: Google earth V 6.2.2.6613. (November 14, 2018). Mandurah, Western Australia. 32◦31’23.59’’S 115◦42’43.34’’E, DigitalGlobe 2018. http://www.earth.google.com [April 4. 2019].

18

2.3 Candidate shellfish species selection

A bioclimatic modelling approach, also known as envelope models, predicts the geographic ranges of organisms as a function of climate, in this case it was used to identify the most appropriate candidate species for shellfish reef restoration based on the climate of the Peel-Harvey Estuary (Araújo & Peterson, 2012). This involved determining the monthly extremes in water temperatures and salinities throughout the year and published literature on the environmental tolerances of the five aforementioned reef-forming oyster species.

Water temperature and salinity data, recorded monthly from two sites in each of the Peel

Inlet and in the Harvey Estuary, were provided by the Western Australian Department of

Water Regulation (Department of Water and Environmental Regulation, 2019). The sample stations were chosen as they represented a continuous time series of data from

2002 to 2016 and were distributed evenly throughout the Peel Inlet and Harvey Estuary.

This data was used to calculate the monthly extremes of the range in temperatures and salinities throughout the year, taken as the 90th and 10th percentiles.

The monthly extremes in water temperature and salinity were then used to determine the time each of the candidate oyster species spent outside their physiological tolerance range. To validate this approach the top two species, S. glomerata and O. angasi, were then underwent tank trials in which they were subjected, for ten weeks, to the extreme winter, taken as the approximate mean of 10th percentile of water temperature and salinity during the winter months (June to August) and extreme summer conditions, taken as the approximate mean of 90th percentile of each variable during the summer months

(December to February).

2.4 Laboratory experiments

Saccostrea glomerata and O. angasi were obtained from Blue Lagoon Mussel Farm in

Cockburn Sound and Albany Shellfish Hatchery in Princess Royal Harbour, respectively,

19

during October/November 2018 and transported to the laboratory at the Freshwater Fish

Group and Fish Health Unit (Murdoch University). There were no mortalities during the transportation from the collection site to the laboratory. On arrival, the oysters were placed into a holding tank that was previously set up to replicate the approximate temperature and salinity of marine waters in which they were growing, at that time of collection (i.e. ~17◦C and 35ppt; Fig. 2.6).

The tank experiments comprised of 12 tanks each of which contained 40 L of water, a wave maker, biological filter, aerator and, in the case of the summer condition tanks, a heater. Ten small (20-30 mm) and four large (50-70 mm) O. angasi and S. glomerata were randomly selected and placed in each tank. The large individuals had valvometers attached to monitor their valve gaping behaviour (see later), while the smaller individuals were used to measure survival and body condition (Fig. 2.6).

Over the course of a week, water temperatures and salinities in the four summer, four winter and four marine (control) treatment tanks were acclimatised to the required temperatures and salinities. In the case of the summer treatment tanks, salinity was increased by ~2ppt and temperature by one Celsius degree each day, to reach 26◦C and

48ppt. In the case of the winter treatment tanks, salinity was decreased by ~2ppt and temperature by ~0.5◦C to reach 15◦C and 14ppt (Table 2.1; Fig. 2.6).

Data loggers were used to record the water temperature in each tank continuously, whereas salinity, dissolved oxygen and ammonia concentrations were monitored daily and calcium and pH weekly. Each tank was under constant aeration ensuring that dissolved oxygen concentrations remained in excess of 8 mg/L. A water conditioner

(Seachem Prime) was added to the water, and 25% water changes were undertaken biweekly, except when ammonia levels exceeded 0.5 or 1 mg/L, in which case a 50% or

100% water change, respectively, was conducted immediately. Ammonia was measured

20

using the API (Aquarium Pharmaceuticals Industries) Ammonia test kit and the respective salinities and pH of those tanks.

Table 2.1. The water temperature and salinities for each condition based on the monthly percentiles of the Peel-Harvey Estuary.

Condition Temperature (°C) Salinity (ppt)

Marine (control) 17 35 Summer 25 48 Winter 15 14

2.4.1 Feeding

Oysters were fed Instant Algae Shellfish Diet 1800 (Reed Marciulture). This diet comprised of six marine microalgae i.e. Isochrysis sp., Pavlova sp., Tetraselmis sp.,

Chaetoceros calcitrans, Thalassiosira weissflogii and Thalassiosira pseudonana. The shellfish diet was used due to its nutritional value, long shelf life and easy administration being a liquid.

The quantity of feed administrated was determined by formula provided by the manufacturer, i.e. Shellfish Diet feed (ml/d) = oyster meat (dry mass, g) x 0.335, where the average dry mass of individual small and large O. angasi were 0.07 g and 0.07 g, respectively and those for S. glomerata were 0.031 g and 0.200 g, respectively. The

O. angasi oysters did not differ in size, the small and large oysters were given the same mass of feed. Thus, the recommended dosage of Shellfish Diet for each tank at the beginning of the experiment was 2 ml/day. However, as this formula is used for calculating food ration for increasing growth performance of shellfish, Reed Mariculture recommended that the oysters in our trials be fed this amount twice per week. Thus, the oysters were fed 1 ml two times per week, with the feed administered using a 3 ml syringe in front of a wave maker to ensure distribution throughout the tank. Adjustments to the feeding amount were adjusted due to the mortality of oysters.

21

2.4.2 Survival, body condition and valve activity analyses

The number of individuals of each species in each tank were recorded daily at a consistent time (i.e. 9 am). Any individuals that had died, i.e. those where the abductor mussel had relaxed facilitating an extra wide valve gape, were removed from the tank.

At the end of the experiment comparisons of the oyster survival between species and conditions was undertaken using the packages ‘survival’ (Therneau, 2015) and

‘survminer’ (Kassambara et al., 2017) in R (R Core Team, 2017), in which a non- parametric statistic (the Kaplan-Meier estimator) is used to estimate the probability of survival over time, assuming a binomial distribution. Pairwise log-rank test was then used to compare the survival curves of the different species/treatments and the resultant P- values derived through a chi-squared distribution, in which a P-value <0.05 is considered statistically significant, whereas a P-value >0.05 indicates that, from the data, there is no evidence that they differ significantly different. A cox proportional hazard model (coxph) was also conducted for further analysis in R (Cox, 1972). The hazard function determines the probability of an event or its hazard (survival) if the oyster survived up to a particular time point (end of the trial). The hazard function considers covariates when comparing the survival of groups, in this case the hazard ratios considers the survival of the two oyster species in each treatment. A hazard ratio greater than 1 indicates an increased risk of death, and a hazard ratio less than 1 indicates a decreased risk of death.

To calculate body condition index (BCI), a random sample of 20 oysters of each species was euthanised prior to the treatments and the shell lengths measured using numeric callipers to the nearest 0.1 mm. The digestive gland and soft tissue were removed from the shell and dried in an oven for 96 h at 60oC (Aguirre-Velarde et al., 2018). The dry weight of the tissue was weighed to the nearest 0.01 g. This process was repeated at the end of the experiment or, in some cases, following mortality during the experiment.

The body condition index was calculated for each individual oyster by the following equation (Riascos et al., 2012): dry weight (g) BCI= shell length (mm) × 100 22

Analysis of Variance (ANOVA) was then used to determine whether there were any significant differences in the BCI before or among treatments (P<0.05). To determine whether the change in BCI differed between species the change in BCI of each oyster ̂ was calculated as ∆푀푗 = 푀푡푗 − 푀푏, where ∆푀푗 is the change in BCI of the j’th oyster, 푀푡푗 is the BCI of the j’th oyster after treatment and 푀̂푏 is the average BCI before treatment.

The valve activity of four S. glomerata and four O. angasi in each tank were recorded approximately every 10 seconds for the duration of the experiment. Each valvometer comprises of a sensor and a magnet attached to each opposing valve (García-March et al., 2008; Gnyubkin, 2010). The sensor is wired to a CPU, which enables the passing of the data to an SD card. The sensor measures the strength of the magnetic field, based on the distance between the sensor and the magnet, i.e. between the two valves, thereby determining when the oyster’s valves are open or closed or transitioning, i.e. opening or closing (Fig. 2.4).

Fig. 2.4 A sample of the valvometry of an oyster’s activity. Each point represents the position of the oyster every ~10 seconds. When there is a high magnetic field (i.e. 555) reading the state of the oyster is closed due the magnet being closer to the sensor (i.e. 530), in contrast, when the magnetic field is low, the oyster in an open state. Between open and closed states, the oyster is transitioning.

23

To facilitate statistical comparisons between conditions/species the proportion of time the valves of each oyster spent open was calculated. The resultant data was then arcsine square root transformed, the standard approach used for pre-treating proportional data

(Warton & Hui, 2011) and subjected to two-way ANOVA in R, in which condition and species were the factors. A post-hoc test (LSD) was conducted to determine the significant differences between each condition of S. glomerata. The resultant means (and

95% confidence intervals) were then back transformed and plotted.

Prior to analyses, a Pearson’s correlation test was used to confirm that the valve activity of the oysters did not change throughout the duration of the experiment (P>0.05), thus demonstrating that it was appropriate to compare the valve activity of the oysters in the summer condition tanks, which survived for <3 weeks, to those in other conditions, many of which survived the two month study period (r = -0.2795, P = 0.50, Fig. 2.5).

1

0.8

0.6

0.4

0.2 Proportion of time spent of time openProportion

0 0 1 2 3 4 5 6 7 8 Weeks of trial Fig. 2.5 Pearson correlation test between the proportion of time spent open at weekly intervals of the trial.

24

Fig. 2.6 Flowchart illustrating the experimental process.

25

Chapter 3: Results

3.1 Historical and current distribution of reef forming shellfish in the Peel- Harvey Estuary

No historic records of oyster reefs in the Peel-Harvey Estuary were found during the extensive reviews and discussions with local stakeholder groups. However, in the

Dudley Park Canals (bottom left of Fig. 3.1), multiple live specimens of Ostrea stentina

(Aquatic Life Industries pers. comm.) and S. glomerata were found submerged on the limestone walls and on the wooden jetty pylons. The oysters on the limestone wall were dispersed apart from one another rather than aggregated together. However, on the wooden pylons, single oysters were found aggregated with Mytilus galloprovincialis (Blue

Mussels). On the traffic bridge across from the Dudley Park Canals (bottom middle of

Fig. 3.1), oysters were found attached on the underside of the cement pylons. The oysters were found sparsely between the dense aggregations of M. galloprovincialis.

The distribution of oysters throughout the Old Coast Road canals was minimal. When entering the canals there was evidence of both living and dead oyster shells, along the limestone wall and the rocks. Oysters were found individually, submerged, and a few dead shells unsubmerged were observed. Several large alive and smaller dead oysters were found on the cement pylons of the traffic bridge. Heading east into the canals, a rock bed was found, with evidence of alive and dead oysters. Heading north in the canals, another rock bed only had evidence of dead oyster shells.

In the northern canals, there was evidence of live oysters on wooden jetty pylons, and live and dead oyster shells along a limestone wall. Parallel to Fairbridge Road, there are several jetties, in which there was evidence of individual live and dead oyster shells. This was also the case along the north eastern limestone wall of the marina (top right hand of

Fig. 3.1). Similarly, to the other sites, the oysters were only found individually or within aggregations of M. galloprovincialis, however, never were there dense aggregations of

26

oysters. Although the distribution of oysters was lacking in abundance, the evidence of live oysters distributed throughout the Mandurah Channel indicates that oysters can survive in at least the more marine areas of the estuary subjected to daily tidal exchange.

Conversations with the Mandurah Community Museum, Fisherman’s association and commercial fishers, supported the field survey, in that there is no evidence to support historical or present oyster reefs in the Peel Harvey Estuary.

27

Fig. 3.1 Locations of observed oyster, within the Peel Harvey Estuary. where green represents live oysters, orange represents live and dead oysters and red represents dead oysters. Taken from: Google earth V 6.2.2.6613. (November 14, 2018). Mandurah, Western Australia 32◦31’22.59’’S 115◦42’43.34’’E,. DigitalGlobe 2018. http://www.earth.google.com [April 4. 2019].

28

3.2 Environmental parameters & species selection

Median water temperatures derived from data recorded between March 2002 and May

2018 at four sites in the Peel-Harvey Estuary ranged from 12◦C in July to 24.6◦C in

January (Department of Water and Environmental Regulation, 2019). Median salinities of the Peel-Harvey Estuary, recorded for the same period at those sites, typically ranged from 22ppt in August to 44ppt in March (Fig. 3.2).

The five potential candidate oyster species were ranked based on the minimum percentage time exposed outside their tolerance range based on the environmental envelope of each species and the monthly 10th and 90th percentile of temperature and salinity in the Peel-Harvey Estuary (Fig. 3.2). The upper temperature tolerances of the five oyster species are under extreme high temperatures exhibited in the Peel-Harvey

Estuary. However, the low temperature tolerance of S. cucullata, S. scyphophilla and

S. echinata, exposes them to the extreme low temperatures that occur in the Peel-

Harvey Estuary (Fig. 3.3; Table 3.1). In terms of salinity, all of the species’ lower salinity tolerances expose them to the extreme low conditions in the Peel Harvey Estuary

(Fig. 3.3). The least time exposed is S. glomerata, followed by S. cucullata, O. angasi,

S. scyphophilla and S. echinata (Table 3.1). The upper salinity tolerances of all species except S. glomerata are lower than the extreme summer salinity recorded, in the Peel-

Harvey Estuary. Based on the percentage of exposure to the 10th and 90th percentile of winter and summer temperature and salinities, S. glomerata and O. angasi were the most suitable of the five species (Table 3.1).

To validate the above bioclimatic modelling approach, S. glomerata and O. angasi were then subjected to tank trials in which the extremes in water temperatures and salinities, i.e. the approximate mean of the 10th percentile during the winter months (June, July and

August), i.e. 15◦C and 14, respectively, and that of the 90th percentiles during the summer months (December, January and February), i.e. and 26◦C and 48, respectively.

29

Fig. 3.2 The monthly temperature and salinity recorded in the Peel Harvey Estuary between 2002- 2016 (Department of Water and Environmental Regulation, 2019). The box plots represent the 25th, 75th percentiles with the mean in the middle, and the whiskers represent the 10th and 90th percentiles.

Table 3.1 The predicted percentage of time exposed by five oyster species outside their tolerance range throughout the year.

Species Temperature Salinity Reference Saccostrea glomerata 0% 8% (Nell & Holliday, 1988) Ostrea angasi 0% 25% (Nell & Gibbs, 1986) Saccostrea cucullata 25% 50% (Venture, 2016) Saccostrea scyphophilla 25% 75% (Venture, 2016) Saccostrea echinata 42% 83% (Venture, 2016)

30

Fig. 3.3. The minimum (blue) and maximum (red) temperature and salinity tolerances of species, compared to the lower and upper monthly (10th & 90th percentiles) temperature and salinity (black) exhibited in the Peel-Harvey Estuary (Department of Water and Environmental Regulation, 2019).

31

3.3 Laboratory Trial

3.3.1 Survival analyses

In the marine condition, only 3 of the 56 O. angasi died over the 62-day period of the trial. The first two mortalities occurred on day 25 and 26 of the trial and the final mortality on day 37, resulting in 53 oysters at the end of the trial survival of O. angasi (Fig. 3.4).

In summer conditions the first O. angasi mortality occurred after day four of the trial, with none of the 56 oysters remaining alive at the end of the trial. Mortality in the winter condition, started 16 days into the treatment, leaving only 16 oysters remaining. The probability of survival of O. angasi were 0.946, 0.189 and 0, for marine, winter and summer conditions, respectively (Fig. 3.4).

Saccostrea glomerata, suffered few mortalities in the marine and winter conditions with

52 oysters and 55 oysters remaining in their respective treatments after the trial. In contrast, the summer condition caused high mortalities, leaving only a single oyster alive at the end of the trial period. The probabilities of survival for S. glomerata were 0.946,

0.982 and 0.036, for marine, winter and summer conditions, respectively (Fig. 3.4).

Overall, there was a statistically significant difference between the probability of survival of O. angasi and S. glomerata between the treatments (Table 3.2), and a difference between the two species within each treatment (Table 3.3).

The marine condition serves as a reference to calculate the hazard ratio for summer and winter conditions for S. glomerata and O. angasi. For O. angasi, the summer condition has a hazard ratio of 240, and 22 for the winter conditions, both are statistically significantly different from the marine conditions, demonstrating a higher risk of death within the two conditions (P<0.001; Fig. 3.5). For S. glomerata, the summer condition has hazard ratio of 62.65 and in the winter condition, 0.33. As shown by the forest plot, the respective 95% confidence intervals, the summer condition hazard ratio is statistically significant different to the marine condition (P <0.001) while the winter condition is not

(P >0.05; Fig. 3.5). The results thus demonstrate statistically that, for both species, there

32

is a low probability of survival in summer conditions and, for O. angasi also in winter conditions.

Fig. 3.4 Mean and 95% confidence intervals of the probability of survival of Ostrea angasi and Saccostrea glomerata in experimental treatments; marine, summer and winter conditions.

33

Table. 3.2 P-values derived from a ‘survdiff’ test in R, for the survival of the oyster species; Ostrea angasi and Saccostrea glomerata, between treatments; marine, summer and winter conditions.

Species Treatment Treatment Sig. (p-value)

O. angasi Marine Summer <0.001 Winter <0.001

Summer Winter <0.001

S. glomerata Marine Summer <0.001 Winter >0.05

Summer Winter <0.001

Table. 3.3 P-values derived from a ‘survdiff’ test in R, comparing the oyster species; Ostrea angasi and Saccostrea glomerata, survivability within treatments; marine summer and winter conditions.

Species Treatment Sig. (p-value) O. angasi S. glomerata Marine >0.05 Summer <0.001 Winter <0.001

34

a)

b)

Fig. 3.5 Forest plots illustrating the hazard ratio of a) Ostrea angasi and b) Saccostrea glomerata in the experimental treatments; summer and winter conditions compared to the marine condition.

35

3.3.2 Valve Activity

The highest proportion of time spent by O. angasi and S. glomerata with valves open was observed in the marine condition (0.53, 0.43), followed by the winter condition (0.26,

0.15) and lastly the summer condition (0.04, 0.11; Fig. 3.6). No statistically significant difference was found between the two species within each of the three conditions

(P>0.05). Individually, across the three conditions there was a statistically significant difference between the time spent open for O. angasi (P<0.05), in contrast, there was no statistically significant difference between the time spent open for S. glomerata (P=0.05).

However, post-hoc test, showed that there was a statistically significant difference between the time spent open, between S. glomerata in the marine and winter conditions

(P<0.05), and marine and summer conditions (P<0.05).

Fig. 3.6 The proportion of time, Ostrea angasi and Saccostrea glomerata spent open within the respective treatments; marine, summer and winter conditions.

36

3.3.3 Body Condition Index

The body condition index (BCI) of O. angasi did not change significantly from 0.19 following the subjection to the marine (0.186), summer (0.19) and winter conditions (0.16;

P>0.5; Fig. 3.7). As for S. glomerata, there was an increase in BCI in the marine and winter condition (0.14, 0.13) and decrease in BCI in summer conditions (0.094; Fig. 3.7), however, the differences were not statistically significant (P>0.05).

There was a statistically significant difference when comparing the two species within the treatments (P<0.001; Fig. 3.8). Pairwise comparisons showed no statistically significant difference in the change of BCI between the two species in the marine or summer conditions (P>0.05), however there was a statistically significant difference in the winter condition (P<0.001; Fig. 3.8). In the winter condition, O. angasi observed a decrease in BCI by 0.035, in contrast, S. glomerata observed an increase in BCI by 0.037

(Fig. 3.8).

Fig. 3.7 The body condition index (BCI) of Ostrea angasi and Saccostrea glomerata, comparing the BCI before the trial to the experimental treatments; marine, summer and winter conditions.

37

Fig. 3.8 The body condition index (BCI) of Ostrea angasi and Saccostrea glomerata, comparing the change in BCI in the experimental treatments; marine, summer and winter conditions.

38

Chapter 4: Discussion

Through the exploration of physiological tolerances of five Australian oyster species and the extremes in environmental conditions in the Peel-Harvey Estuary this study identified the Australian Flat Oyster Ostrea angasi and the Sydney Rock Oyster Saccostrea glomerata as the most suitable candidates for oyster reef restoration in the Peel-Harvey

Estuary. The study then aimed to validate this bioclimatic modelling approach through laboratory experiments to explore survivability and physiological (behavioural and biological) responses to the absolute extreme conditions that occur in water temperature and salinity during winter and summer in the estuary, thereby providing support that these species could be used in almost any location throughout the system for restoration purposes.

4.1 Distribution of oysters in the Peel-Harvey Estuary

Although the information on the historic and current distribution of reef-forming shellfish in south-western Australia is very limited, the lack of any evidence of oyster reefs in the

Peel-Harvey Estuary suggests that such reefs have never existed in this system. This conclusion is based on museum records, which, unlike estuaries elsewhere in Australia, there is no record of oysters providing food or lime products for early settlers.

Furthermore, this conclusion is supported by the fact that the system is almost completely devoid of hard structure and therefore does not provide habitat conducive for oyster spat settlement. However, there is anecdotal evidence that beds of Blue Mussels

(Mytilus galloprovincialis) existed in recent decades in the southern Peel Inlet (Mandurah

Licensed Fishermen’s Association, pers. comm.).

Given the current distribution of O. angasi and S. glomerata, it was not surprising that these two species were the most suitable candidates, based on their temperature and salinity tolerances, for oyster reef restoration in the Peel Harvey Estuary. Furthermore, the extensive distribution of S. glomerata around Australia, including many degraded estuaries on the east coast is indicative of its ability to tolerate and flourish under a wide range of environmental conditions. For example, even in the heavily urbanised Sydney 39

Estuary, NSW, the population of S. glomerata, which collapsed during the 1980s returned ~25 years later and this species is now highly abundant throughout the entire estuary, noting that the temperature and salinity ranges in that estuary were 16.8 to

23.8◦C and 0-35ppt, which differ markedly to the extreme conditions in of the Peel-

Harvey Estuary (12-25◦C and 14-48ppt). The ability for S. glomerata to tolerate and perform in such a wide range of environmental conditions has led to extensive commercial aquaculture of this species outside of its natural range, such as in Cockburn

Sound, 40 km to the north of the Peel-Harvey Estuary.

In contrast, to the information available for S. glomerata that demonstrates that it is often found in temperatures that approach those of the extreme summer conditions in the Peel-

Harvey Estuary, few records of O. angasi in coastal waters north of the Peel-Harvey

Estuary exist, suggesting that this species is close to its northern most limit for temperature in ocean waters at this latitude. In Port Phillip Bay (Cole et al., 2016) and

Oyster Harbour (WA), O. angasi was the species of choice for restoration by The Nature

Conservancy. The species was chosen due to the temperature and salinity of those locations falling well within the species physiological tolerable range and suitability is further supported by the presence of live O. angasi throughout the bay as well as evidence of historical reefs. Oyster reefs of O. angasi were also found to be the most lost reefs in Australia with only one reef still remaining in Georges Bay, Tasmania, flourishing in water temperatures ranging between 10 to 18◦C and salinities 10 to 35ppt

(Mitchell et al., 2000; Crawford & White, 2019).

40

4.2 Survivability and physiological responses under different environmental conditions

Temperature and salinities in temperate estuaries are inherently highly variable and consequently expose oysters to conditions that are outside their optimal limits for extended periods of time, including high and low temperature and salinities, very little to no oxygen and rapid salinity changes (Heilmayer et al., 2008; Lowe et al., 2017). This can have direct impact on all physiological processes, such as respiration, filtration, metabolism, and feeding, that are necessary for the growth and survival of oysters

(Heilmayer et al., 2008). As sessile organisms with no motility, oysters cannot avoid unfavourable environmental conditions, therefore in order to survive, oysters have developed a behavioural adaptation, i.e. valve closure. (Akberali & Trueman, 1985;

Riisgård et al., 2006; Rodland et al., 2008; Redmond et al., 2017). The open/closed valve state of oysters influences all physiological process, such as filtration, metabolism, respiration and reproduction occur (Riisgård et al., 2006; Redmond et al., 2017), but can also minimise death during short periods of unfavourable environmental conditions

(Hoyaux et al., 1976; Akberali & Trueman, 1985; Lowe et al., 2017). For example, in estuaries, valve closure can provide a period of grace to differing salinities, which occur following freshwater discharge, that can cause changes to the oyster’s osmotic concentration that can consequently cause osmotic shock and death. However, when an oyster’s valves are closed for extended periods of time, aerobic scope decreases due to a lack of oxygen supply to their tissues causing a change to anaerobic metabolism that can also result in stress and eventual death (Loosanoff, 1942; Akberali & Trueman,

1985). Furthermore, as temperatures increase, the metabolic rate and thus the need for oxygen likewise increases, which requires exposing their gills to the outside environment, even although this can be detrimental.

The high salinities and temperatures associated with the extreme summer conditions in the Peel-Harvey Estuary had a detrimental effect on the survival both oyster species, with no O. angasi and only one S. glomerata surviving the 62-day trial. In the case of O.

41

angasi, this species was ~3ppt outside its upper tolerance range for salinity of 45ppt, but well within its upper thermal tolerance limit of 29oC. Furthermore, mortality began seven days following exposure and almost all were deceased by the end of the first week. The low probability of survival within the summer conditions of O. angasi is consistent with

Nell and Gibbs (1986), who found, when O. angasi was exposed to salinties outside its tolerance range, total mortality occurred after eight days. Similarly, in the winter conditions, survival of O. angasi, which was 6ppt outside its lower salinity tolerance limit, was likewise low. Mortality in winter conditions, however, began after ~two weeks and

16 (out of 56) survived to the end of the trial. The greater survival during winter than in summer, despite being further outside its tolerence range, probably reflects the lower metabolic activity and thus reduced physiological processes that occurs when temperatures are lower.

Similarly, the survial of S. glomerata was low in summer conditions. However, unlike O. angasi, S. glomerata was well within its salinity (and temperature) tolerence range. It is thus likely that the low survival of S. glomerata in the summer conditions was related to the combination of both relatively high temperatures and salinity, which is consistent with the many studies that have documented the deleterious synergistic effects that temperature and salinity can have to oysters (Heilmayer et al., 2008; Rybovich et al.,

2016; Casas et al., 2018a). For example, Lowe et al. (2017) and La Peyre et al. (2015), quantified high rates of mortalities in high temperatures and high salinities. The study by

Lowe et al. (2017) demonstrated that mortality increased with increasing temperatures, particularly when salinities were likewise high. Thus, the highest rates of mortality observed when temperatures exceeded 30°C and salinities >15ppt.

The differences in survival of the oystes were reflected by significant differences in behaviour, both O. angasi and S. glomerata spent a greater time open in the marine conditions (0.53 and 0.43, respectively), compared to that in the summer and winter for

O. angasi (0.04 and 0.26, respectively) and summer for S. glomerata (0.1), when survival was least. A lowered valve due to environmental conditions (temperature and salinity)

42

outside of the tolerable range decreases the amount of time oysters spent open. Studies conducted by Casas et al. (2018a,b) revealed similar results using oyster species, C. virginica. A greater proportion of time open within optimum salinities (6-25 ppt) and temperature (20°C), than that of conditions outside of the oyster’s optimum range (3ppt,

10C and 30°C).

The low survival and valve openness of S. glomerata in summer conditions was also consistent with the BCI for this species in summer conditions, an index which reflects the health of an organism (Mason & Nell, 1995; Heilmayer et al., 2008; Riascos et al., 2012;

Verdelhos et al., 2015; Taylor et al., 2017). The relationship between temperature and

BCI was consistent with the results of a study by Lowe et al. (2017), who revealed BCI of C. virginica increased as temperatures decreased, with the highest mean BCI occurring when exposed to temperatures below 17.5°C, whereas a low mean BCI was recorded at temperatures equal to or greater than 25°C.

In the case of O. angasi, caution must be exercised when interpreting the results as the extremely high mortality in the first week in summer conditions did not produce comparable BCI values. However, the lowered BCI in O. angasi exposed to winter conditions suggests that the were stressed and not favourable to the conditions, which was consistent with this species being outside its salinity tolerance. These results may explain a correlation between an oyster’s valve activity and its body condition index.

With a greater time spent open, oysters tend to have a greater body condition index, in contrast with a lowered time open, there is a decrease in body condition index. This is due to the association between valve openness and the ability to filter water that provides food and oxygen to the animals’ tissues in order to grow.

43

4.3 The ecological and social risk of oyster reef restoration in the Peel- Harvey Estuary

Oyster reefs provide a range of highly valuable ecosystem services that can potentially benefit the Peel-Harvey Estuary (Coen et al., 2007). When implementing such restoration projects, it is essential, however, to identify the ecological and social risks that may occur by carrying out the project (Menzie et al., 2013; Methratta et al., 2013), especially since the Peel-Harvey Estuary is listed as a Wetland of International Important under the Ramsar Convention on Wetlands and no oyster reefs historically or currently exist in this system (Hale & Butcher, 2007). Oyster reefs can have direct or indirect interactions with all biota groups within the estuary and can have positive and negative impacts on water quality, other benthic communities, submerged aquatic vegetation

(SAVs), such as , soft-bottom invertebrate communities, planktonic communities (phytoplankton and zooplankton), pelagic communities (Coen et al., 2007;

Richkus & Menzie, 2013; Johnston et al., 2015) as well as birds, mammals and fringing vegetation. In particular, greater considerations are required in the case of S. glomerata as this species is not native to south-western Australia, but is used in aquaculture throughout this region. The introduction of oysters, could also potentially introduce diseases, such as QX, Bonamiosis, and/or invasive species into the estuary (Nell, 2001).

Furthermore, in the case of S. glomerata, a comprehensive risk assessment is required as this species is not native to the region and there has not been a very good track record historically for introducing new species to an area (Ruesink et al., 2005).

The Peel-Harvey Estuary host highly valued commercial and recreational fisheries for

Blue Swimmer Crabs (Portunus armatus), and finifish, such as Sea Mullet (Mugil cephalus), Cobbler (Cnidoglanis macrocephalus) and Yellow-fin Whiting (Sillago schomburgkii; Johnston et al., 2015). With 11 licensed fisheries operating in the estuary and a vast number of recreational users, catches of both P. armatus and finfish typically exceed 200 tonnes per year (Johnston et al., 2017; Gaughan & Santoro, 2018). The commercial catches of finfish listed above are the greatest in the Peel-Harvey Estuary

44

than other regions within the whole of Australia’s West Coast bioregion (Johnston et al.

2017). The increase of habitat, prey availability and shelter provided by oyster reefs can greatly benefit both the ecology and the social uses of the estuary in the Peel Harvey estuary. Although the restoration of oyster reefs in the Peel-Harvey Estuary could highly benefit commercial and recreational fishing sectors, some changes might be unwelcoming to the users of the Estuary. With almost 10% of the Mandurah population owning boats, boating channels may need to change to navigate around the reefs in shallow waters (Valesini et al., 2019).

Several Aboriginal heritage sites are also located around the Peel-Harvey Estuary which are protected by the Western Australian Aboriginal Heritage Act (1972) including campsites at the Serpentine River mouth and Island Point, and a ceremonial site at Egg

Island (Hale and Butcher 2007). The restoration of oyster reefs in the estuary will require liaison with the local aboriginal community in development of the reef placement and also the management of the reefs.

4.4 The potential of restoration in the Peel-Harvey Estuary using oyster reefs

Oysters also aid in nitrogen removal through the process of denitrification, that returns nitrogen into the atmosphere as an inert gas and by sequestrating nitrogen into their shells, tissues and biodeposits (Kellogg et al., 2013; Hoellein & Zarnoch, 2014). For example, in the Mission-Aransas Estuary, the 18.11 km2 oyster reef with an average density of 100 live oysters m2 were estimated to remove a total of 9100 kg N (502kg km-

2 N) via coupled denitrification of biodeposits and up to 4550 kg N (251.3kg km-2 N) via the burial of biodeposits in sediments per annum (Pollack et al., 2013). The ability to improve water quality has made a key goal for using oyster reefs for the restoration of degraded estuaries around the world (Carmichael et al., 2012; Caffrey et al., 2016;

Chakraborty, 2017). In the nitrogen rich Peel-Harvey Estuary that receives high nutrient loading (an annual 500 tonnes of TN) from nearby agricultural, industrial run off and urbanised areas, the process of denitrification can help to counteract eutrophication

45

(Valesini et al., 2019). Oyster reef restoration in other parts of the world have also demonstrated that such restoration efforts can lead to substantial increases in fisheries production. In one the world’s largest estuaries (11,601 km2), Chesapeake Bay, USA, restoration efforts were accompanied by a ten million dollar increases in annual commercial catches of Blue Crabs (Callinectes sapidus). While the restoration only restored a fraction of the oyster reefs that once had the capacity to filter its entire volume of the estuary (71.5 GL) in ~three days (Newell, 1988; Coen & Luckenbach, 2000), it was estimated that production of fish and crabs by oyster reefs was 0.26 kg m-2 y-1 for the lifetime of the reef (Peterson et al., 2003). Since oyster reefs are biogenic and self- sustaining, a reef surviving for 40 years can augment an accumulative amount of 10 kg m-2 of fish and crustaceans (Peterson et al., 2003).

Although the area (131 km2) and volume (~150GL) of the Peel Harvey Estuary is much less than that of Mission-Aransas Estuary and Chesapeake Bay, it is still the largest estuary in south-western Australia, and most of the estuary is ~ 1m deep and characterised by muddy/silty and unstructured floor beds. Thus, a restoration goal of

0.5% of the estuary area (0.67 km2), stocked with ~45 million adult oysters (50 oysters m2) would facilitate the removal of ~200 kg of N and filter the entire volume in ~20 days, assuming that the reefs were distributed in a way that facilitates the movement of the water over the area and a filtration rate of the Crassostrea virginica

(Newell, 1988). This time is relatively fast given the residence time of water due to seasonal flushing and tidal amplitude in the estuary is 22 – 55 days (Valesini et al., 2019).

46

Chapter 5: Conclusions

5.1 Future recommendations for oyster reef restoration in the Peel-Harvey Estuary.

While this study has demonstrated that there is potential for shellfish reef restoration in the Peel-Harvey Estuary, further modelling should be undertaken that considers the spatio-temporal nature of, in particularly, temperature and salinity concentrations, such as through the application of a habitat suitability index (HSI) model (Paolisso & Dery,

2010). Deciding where to construct oyster reefs will also require further research to understand the ecological and social risks. Successful and sustainable oyster reef restoration efforts require sites that will support long-term growth and the survival of oysters, i.e. fall within their physiological thresholds. Field trials could be useful to support the current study, and would allow for the effects of dissolved oxygen changes, phytoplankton abundances in the Peel-Harvey Estuary to be assessed. If reefs were constructed to facilitate filtration and improve water quality, for example, locations near to either of the entrance channels (Mandurah and Dawesville Cut) would not be ideal because the oysters would filter water undergoing tidal exchange. Instead it would be better, providing water quality was suitable, to locate the reefs in areas with high residence time that often harbour high density of phytoplankton. Several different sized reefs located throughout the Peel Inlet and Harvey Estuary would be most beneficial, not only increasing water quality, water turnover and flushing, but also habitat complexity of the estuary. Restoration efforts should liaise with a range of stakeholders, including the local communities, indigenous groups and local councils for the placement and installation and monitoring of the reefs, encouraging community engagement and education opportunities. Fisheries and Department of Transport should be liaised with to ensure the placement of the reefs will not impose hazards to other estuarine users.

47

5.2 Concluding remarks

Oyster reefs are one of the most threatened marine habitats in the world, and in particularly, Australia. In addition to the loss of oyster reefs, the valuable ecosystem services that they provide (water filtration and regulation, habitat provision, species production, shoreline protection) have also been lost. Such ecosystem services would be highly valuable to degraded microtidal estuaries as well as helping to alleviate some of the effects of sea level rise brought about by climate change. Although reefs of O. angasi and S. glomerata have not historically occur in the Peel-Harvey Estuary, both species appeared to occur in small numbers on the hardened shorelines of the estuary channel, indicating that at least they can survive in the more marine parts of the estuary.

Laboratory results demonstrated that O. angasi would not withstand long-term exposure to the extremes in temperatures and salinities associated with summer and winter in the basins of the Peel-Harvey Estuary, as indicated by their survival and, to a lesser extent, valve activity and lowered body condition. However, S. glomerata had a much greater probability of survival in winter, supported by similar valve activity to the control/marine conditions and an increase in body condition, and ~50% survived for 3 weeks following exposure to extreme summer conditions. Given the results of the laboratory studies, and the natural distribution of the two species, if restoration efforts were to be undertaken in the Peel-Harvey Estuary, S. glomerata would be the most suitable candidate, noting that this species is not native to the Peel-Harvey region.

48

References

Aguirre-Velarde, A., Jean, F., Thouzeau, G., & Flye-Sainte-Marie, J. (2018). Feeding behaviour and growth of the Peruvian () under daily cyclic hypoxia conditions. Journal of sea research, 131, 85-94. Akberali, H. B., & Trueman, E. (1985). Effects of environmental stress on marine bivalve molluscs. In Advances in marine biology (Vol. 22, pp. 101-198): Elsevier. Angell, C. L. (1986). The biology and culture of tropical oysters (Vol. 13): WorldFish. Araújo, M. B., & Peterson, A. T. (2012). Uses and misuses of bioclimatic envelope modeling. Ecology, 93(7), 1527-1539. Australia, C. o. (2002). Australian Catchment River and Estuary Assessment 2002. National Land and Water Resources Audit, Canberra. Beck, M. W. (2009). Shellfish reefs at risk: a global analysis of problems and solutions: Nature Conservancy. Beck, M. W., Brumbaugh, R. D., Airoldi, L., Carranza, A., Coen, L. D., Crawford, C., . . . Kay, M. C. (2011). Oyster reefs at risk and recommendations for conservation, restoration, and management. BioScience, 61(2), 107-116. Bower, S. M. (2015). Synopsis of Infectious Diseases and Parasites of Commercially Exploited Shellfish: of Oysters. Retrieved from http://www.dfo-mpo.gc.ca/science/aah-saa/diseases-maladies/bonostoy- eng.html Bower, S. M., & Kleeman, S. N. (2011). Synopsis of Infection Disease and Parasites of Commcercially Exploited Shellfish: Marteilia sydneyi of Oysters. Retrieved from http://www.dfo-mpo.gc.ca/science/aah-saa/diseases maladies/marsydoy-eng.html Brearley, A., & Hodgkin, E. P. (2005). Ernest Hodgkin's Swanland: estuaries and coastal lagoons of Southwestern Australia. Crawley, W.A: University of Western Australia Press. Caffrey, J. M., Hollibaugh, J. T., & Mortazavi, B. (2016). Living oysters and their shells as sites of nitrification and denitrification. Marine Pollution Bulletin, 112(1-2), 86-90. doi:10.1016/j.marpolbul.2016.08.038 Carmichael, R. H., Walton, W., & Clark, H. (2012). Bivalve-enhanced nitrogen removal from coastal estuaries. Canadian Journal of Fisheries and Aquatic Sciences, 69(7), 1131-1149. doi:10.1139/f2012-057 Casas, S. M., Filgueira, R., Lavaud, R., Comeau, L. A., La Peyre, M. K., & La Peyre, J. F. (2018a). Combined effects of temperature and salinity on the physiology of two geographically-distant eastern oyster populations. Journal of Experimental Marine Biology and Ecology, 506, 82-90. doi:https://doi.org/10.1016/j.jembe.2018.06.001 Casas, S. M., Lavaud, R., La Peyre, M. K., Comeau, L. A., Filgueira, R., & La Peyre, J. F. (2018b). Quantifying salinity and season effects on eastern oyster clearance and oxygen consumption rates. Marine Biology, 165(5), 1-13. doi:10.1007/s00227-018-3351-x Chakraborty, P. (2017). Oyster reef restoration in controlling coastal pollution around India: A viewpoint. Marine Pollution Bulletin, 115(1-2), 190-193. doi:10.1016/j.marpolbul.2016.11.059

49

Cloern, J. E. (1982). Does the benthos control phytoplankton biomass in south San Francisco Bay. Marine ecology progress series. Oldendorf, 9(2), 191-202. Coen, L. D., Brumbaugh, R. D., Bushek, D., Grizzle, R., Luckenbach, M. W., Posey, M. H., . . . Tolley, S. G. (2007). Ecosystem services related to oyster restoration. Marine Ecology Progress Series, 341, 303-307. Coen, L. D., & Luckenbach, M. W. (2000). Developing success criteria and goals for evaluating oyster reef restoration: Ecological function or resource exploitation? Ecological Engineering, 15(3), 323-343. doi:10.1016/S0925- 8574(00)00084-7 Cole, V. J., Parker, L. M., O’Connor, S. J., O’Connor, W. A., Scanes, E., Byrne, M., & Ross, P. M. (2016). Effects of multiple climate change stressors: ocean acidification interacts with warming, hyposalinity, and low food supply on the larvae of the brooding flat oyster Ostrea angasi. Marine Biology, 163(5), 125. doi:10.1007/s00227-016-2880-4 Commonwealth of Australia. (2002). Australian catchment, river and estuary assessment 2002. In: National Land and Water Audit, Canberra. Conley, D. J., Kaas, H., Møhlenberg, F., Rasmussen, B., & Windolf, J. (2000). Characteristics of Danish Estuaries. Estuaries, 23(6), 820-837. doi:10.2307/1353000 Cox, D. R. (1972). Regression models and life‐tables. Journal of the Royal Statistical Society: Series B (Methodological), 34(2), 187-202. Crawford, C., & White, C. (2019). Monitoring Framework for Georges Bay Establishment of an integrated water quality monitoring framework for Georges Bay. Dame, R., Dankers, N., Prins, T., Jongsma, H., & Smaal, A. (1991). The influence of mussel beds on nutrients in the Western Wadden Sea and Eastern Scheldt estuaries. Estuaries, 14(2), 130-138. Department of Marine and Harbours. (1992). Dawesville Channel Public Marina and Boat Lauching Facility (C12/92). Retrieved from Department of Primary Industries. (2016). Pacific Oyster Mortality Syndrome (POMS). Retrieved from https://www.dpi.nsw.gov.au/search?query=POMS Department of Water and Environmental Regulation. (2019). Water Information Reporting. Retrieved from http://wir.water.wa.gov.au/Pages/Water- Information-Reporting.aspx Dugas, R., & Roussel, J. (1983). Report on oyster mortalities in Louisiana as a result of excessive freshwater intrusion–1983. Louisiana Department of Wildlife and Fisheries report, 65. Elliott, M., Mander, L., Mazik, K., Simenstad, C., Valesini, F., Whitfield, A., & Wolanski, E. (2016). Ecoengineering with Ecohydrology: Successes and failures in estuarine restoration. Estuarine, Coastal and Shelf Science, 176, 12-35. doi:https://doi.org/10.1016/j.ecss.2016.04.003 García-March, J. R., Solsona, M. Á. S., & García-Carrascosa, A. (2008). Shell gaping behaviour of Pinna nobilis L., 1758: circadian and circalunar rhythms revealed by in situ monitoring. Marine Biology, 153(4), 689-698. Gaughan, D. J., & Santoro, K. (2018). Status Reports of the Fisheries and Aquatic Resources of Western Australia 2016/17: The State of the Fisheries. . Department of Primary Industries and Regional Development, Western Australia.

50

Geraldi, N. R., Simpson, M., Fegley, S. R., Holmlund, P., & Peterson, C. H. (2013). Addition of juvenile oysters fails to enhance oyster reef development in Pamlico Sound. Marine Ecology Progress Series, 480, 119-129. Gillies, C. L., McLeod, I. M., Alleway, H. K., Cook, P., Crawford, C., Creighton, C., . . . Heller-Wagner, G. (2018). Australian shellfish ecosystems: Past distribution, current status and future direction. PloS one, 13(2), e0190914. Gnyubkin, V. (2010). The circadian rhythms of valve movements in the mussel Mytilus galloprovincialis. Russian journal of marine biology, 36(6), 419-428. Grabowski, J. H., & Peterson, C. H. (2007). Restoring oyster reefs to recover ecosystem services. Ecosystem engineers: plants to protists, 4, 281-298. Guo, X., He, Y., Zhang, L., Lelong, C., & Jouaux, A. (2015). Immune and stress responses in oysters with insights on adaptation. Fish & shellfish immunology, 46(1), 107-119. Hale, J., & Butcher, R. (2007). Ecological Character Description of the Peel-Yarlorup Ramsar Site Report. Hallett, C. S., Hobday, A. J., Tweedley, J. R., Thompson, P. A., McMahon, K., & Valesini, F. J. (2018). Observed and predicted impacts of climate change on the estuaries of south-western Australia, a Mediterranean climate region. Regional Environmental Change, 18(5), 1357-1373. doi:10.1007/s10113- 017-1264-8 Harding, J. M., & Mann, R. (2001). Diet and habitat use by bluefish, Pomatomus saltatrix, in a Chesapeake Bay estuary. Environmental Biology of Fishes, 60(4), 401- 409. Haven, D. (1962). Seasonal cycle of condition index oysters in the york and Rappahannock Rivers. Proc. Natl. Shellfish Assocation, 5(1), 42-66. Heilmayer, O., Digialleonardo, J., Qian, L., & Roesijadi, G. (2008). Stress tolerance of a subtropical Crassostrea virginica population to the combined effects of temperature and salinity. Estuarine, Coastal and Shelf Science, 79(1), 179- 185. doi:https://doi.org/10.1016/j.ecss.2008.03.022 Hily, C. (1991). Is the activity of benthic suspension feeders a factor controlling water quality in the Bay of Brest? Marine ecology progress series. Oldendorf, 69(1), 179-188. Hine, P., & Thorne, T. (2000). A survey of some parasites and diseases of several species of bivalve mollusc in northern Western Australia. Diseases of aquatic organisms, 40(1), 67-78. Hine, P. M. (1996). Southern hemisphere mollusc diseases and an overview of associated risk assessment problems. OIE Revue Scientifique et Technique, 15(2), 563-577. doi:10.20506/rst.15.2.940 Hoellein, T. J., & Zarnoch, C. B. (2014). Effect of eastern oysters (Crassostrea virginica) on sediment carbon and nitrogen dynamics in an urban estuary. Ecological Applications, 24(2), 271-286. doi:10.1890/12-1798.1 Hoyaux, J., Gilles, R., & Jeuniaux, C. (1976). Osmoregulation in molluscs of the intertidal zone. Comparative Biochemistry and Physiology Part A: Physiology, 53(4), 361-365. doi:https://doi.org/10.1016/S0300-9629(76)80157-0 Jackson, J. B., Kirby, M. X., Berger, W. H., Bjorndal, K. A., Botsford, L. W., Bourque, B. J., Estes, J. A. (2001). Historical overfishing and the recent collapse of coastal ecosystems. science, 293(5530), 629-637.

51

Johnston, D., Marks, R., & O'Malley, J. (2017). West Coast Blue Swimmer Crab Resource status report 2016. In: Department of Fisheries Western Australia. Johnston, D., Smith, K. A., Brown, J., Travaille, K., Crowe, F., Oliver, R., & Fisher, E. (2015). West Coast Estuarine Managed Fishery (Area 2: Peel-Harvey Estuary) & Peel-Harvey Estuary Blue Swimmer Crab Recreational Fishery: Department of Fisheries. Jupp, T. (2019). 50,000 oyster find new home on Windara Reef. Retrieved from https://www.natureaustralia.org.au/explore/newsroom/50-000-oysters-find- new-home-on-windara-reef/ Kassambara, A., Kosinski, M., Biecek, P., & Fabian, S. (2017). Package 'survminer'. Kellogg, M. L., Cornwell, J. C., Owens, M. S., & Paynter, K. T. (2013). Denitrification and nutrient assimilation on a restored oyster reef. Marine Ecology Progress Series, 480, 1-19. doi:10.3354/meps10331 Kennedy, V. S, Newell, I. R., & Shumway, S. (1996). Natural Environmental Factors. The eastern oyster Crassostrea virginica. Maryland Sea Grant: College Park, 467-513. Kirby, M. X. (2004). Fishing down the coast: historical expansion and collapse of oyster fisheries along continental margins. Proceedings of the National Academy of Sciences of the United States of America, 101(35), 13096-13099. La Peyre, J. F., La Peyre, M. K., Eberline, B. S., & Soniat, T. M. (2013). Differences in extreme low salinity timing and duration differentially affect eastern oyster (Crassostrea virginica) size class growth and mortality in Breton Sound, LA. Estuarine, Coastal and Shelf Science, 135, 146-157. doi:10.1016/j.ecss.2013.10.001 La Peyre, M. K., Serra, K., Joyner, T. A., & Humphries, A. (2015). Assessing shoreline exposure and oyster habitat suitability maximizes potential success for sustainable shoreline protection using restored oyster reefs. PeerJ, 3, e1317. doi:10.7717/peerj.1317 Lenihan, H. S., Peterson, C. H., Byers, J. E., Grabowski, J. H., Thayer, G. W., & Colby, D. R. (2001). Cascading of habitat degradation: oyster reefs invaded by refugee fishes escaping stress. Ecological Applications, 11(3), 764-782. Loosanoff, V. L. (1942). Shell movements of the edible mussel, Mytilus edulis (L.) in relation to temperature. Ecology, 23(2), 231-234. Lowe, M. R., Sehlinger, T., Soniat, T. M., & Peyre, M. K. L. (2017). Interactive Effects of Water Temperature and Salinity on Growth and Mortality of Eastern Oysters, Crassostrea virginica: A Meta-Analysis Using 40 Years of Monitoring Data. Journal of Shellfish Research, 36(3), 683-697. doi:10.2983/035.036.0318 Mann, R. (1978). A comparison of morphometric, biochemical, and physiological indexes of condition in marine bivalve molluscs. Paper presented at the Energy and Environmental Stress in Aquatic Systems, Selected Papers from a Symposium, held at Augusta, Georgia November 2-4, 1977. CONF-771114. p 484-497, 1978. 2 tab, 54 ref. NOAA 04-6-158-44106. Mason, C., J, & Nell, J., A. (1995). Condition index and chemical composition of meats of Sydney rock oysters (Saccostrea commercialis) and Pacific oysters (Crassostrea gigas) at four sites in Port Stephens, NSW. Marine and Freshwater Research, 46(5), 873-881. doi:https://doi.org/10.1071/MF9950873 McComb, A. J. (1995). Eutrophic shallow estuaries and lagoons. Boca Raton: CRC Press.

52

McFarland, K., & Hare, M. P. (2018). Restoring oysters to urban estuaries: Redefining habitat quality for eastern oyster performance near New York City. PloS one, 13(11), e0207368. Menzie, C. A., Salatas, J. H., & Wickwire, T. (2013). Ecological Risks Associated with Oyster Restoration Options for Chesapeake Bay. Human and Ecological Risk Assessment: An International Journal, 19(5), 1204-1233. doi:10.1080/10807039.2013.767111 Methratta, E. T., Menzie, C. A., Wickwire, W. T., & Richkus, W. A. (2013). Evaluating the Risk of Establishing a Self-Sustaining Population of Non-Native Oysters Through Large-Scale Aquaculture in Chesapeake Bay. Human and Ecological Risk Assessment: An International Journal, 19(5), 1234-1252. doi:10.1080/10807039.2013.767112 Meyer, D. L., Townsend, E. C., & Thayer, G. W. (1997). Stabilization and Erosion Control Value of Oyster Cultch for Intertidal Marsh. Restoration Ecology, 5(1), 93-99. doi:10.1046/j.1526-100X.1997.09710.x Mitchell, I. M., Crawford, C. M., & Rushton, M. J. (2000). Flat oyster (Ostrea angasi) growth and survival rates at Georges Bay, Tasmania (Australia). Aquaculture, 191(4), 309-321. doi:https://doi.org/10.1016/S0044- 8486(00)00441-5 Morton, B., & Slack-Smith, S. (2003). First report of the European flat oyster , identified genetically, from Oyster Harbour, Albany, south-western Western Australia. Molluscan Research, 23(3), 199-208. doi:10.1071/MR03005 Mueller, M., & Woodland, H. (2015). Edible Rock Oyster Feasibility Study. Munroe, D., Tabatabai, A., Burt, I., Bushek, D., Powell, E., & Wilkin, J. (2013). Oyster mortality in Delaware Bay: impacts and recovery from Hurricane Irene and Tropical Storm Lee. Estuarine, Coastal and Shelf Science, 135, 209-219. Naturally Resilient Communities. (N.D). Oyster Reefs. Retrieved from http://nrcsolutions.org/oyster-reefs/ Nature Conservancy, T. (2019a). Putting the Oyster Reefs Back Into Oyster Harbour. Restoring Albany's lost shellfish reefs. Retrieved from https://www.natureaustralia.org.au/what-we-do/our-priorities/provide-food- and-water-sustainably/food-and-water-stories/putting-the-oyster-reefs-back- into-oyster-harbour/ Nature Conservancy, T. (2019b). Victoria's Lost Reefs Rediscoverd. Getting shellfish reefs back into Port Phillip Bay. Retrieved from https://www.natureaustralia.org.au/what-we-do/our-priorities/build-healthy- cities/cities-stories/victoria-s-lost-reefs-rediscovered/ Nell, J. A. (2001). The History of in Australia (Vol. 63). Nell, J. A., & Gibbs, P. J. (1986). Salinity tolerance and absorption of L-methionine by some Australian bivalve molluscs. Marine and Freshwater Research, 37(6), 721-727. Nell, J. A., & Holliday, J. E. (1988). Effects of salinity on the growth and survival of Sydney rock oyster (Saccostrea commercialis) and Pacific oyster (Crassostrea gigas) larvae and spat. Aquaculture, 68(1), 39-44. doi:https://doi.org/10.1016/0044-8486(88)90289-X Newell, R. I. E. (1988). Ecological changes in Chesapeake Bay: are they the result of overharvesting the American oyster, Crassostrea virginica. Understanding the estuary: advances in Chesapeake Bay research, 129, 536-546.

53

Ogburn, D. M., White, I., & McPhee, D. P. (2007). The disappearance of oyster reefs from eastern Australian estuaries—impact of colonial settlement or mudworm invasion? Coastal Management, 35(2-3), 271-287. Paling, E. I., van Keulen, M., Wheeler, K., Phillips, J., & Dyhrberg, R. (2001). Mechanical seagrass transplantation in Western Australia. Ecological Engineering, 16(3), 331-339. doi:https://doi.org/10.1016/S0925-8574(00)00119-1 Paolisso, M., & Dery, N. (2010). A Cultural Model Assessment of Oyster Restoration Alternatives for the Chesapeake Bay. Human Organization, 69(2), 169-179. Peterson, C. H., Grabowski, J. H., & Powers, S. P. (2003). Estimated enhancement of fish production resulting from restoring oyster reef habitat: quantitative valuation. Marine Ecology Progress Series, 264, 249-264. Piazza, B. P., Banks, P. D., & La Peyre, M. K. (2005). The potential for created oyster shell reefs as a sustainable shoreline protection strategy in Louisiana. Restoration Ecology, 13(3), 499-506. Pollack, B. J., Cleveland, A., Palmer, T. A., Reisinger, A. S., & Montagna, P. A. (2012). A Restoration Suitability Index Model for the Eastern Oyster (Crassostrea virginica) in the Mission-Aransas Estuary, TX, USA. PloS one, 7(7), e40839. doi:10.1371/journal.pone.0040839 Pollack, J. B., Yoskowitz, D., Kim, H.-C., & Montagna, P. A. (2013). Role and value of nitrogen regulation provided by oysters (Crassostrea virginica) in the Mission-Aransas Estuary, Texas, USA. PloS one, 8(6), e65314. doi:10.1371/journal.pone.0065314 Potter, I. C., Loneragan, N. R., Lenanton, R. C. J., & Chrystal, P. J. (1983). Blue-green algae and fish population changes in a eutrophic estuary. Marine Pollution Bulletin, 14(6), 228-233. doi:10.1016/0025-326X(83)90257-6 Potter, I. C., Veale, L., Tweedley, J. R., & Clarke, K. R. (2016). Decadal changes in the ichthyofauna of a eutrophic estuary following a remedial engineering modification and subsequent environmental shifts. Estuarine, Coastal and Shelf Science, 181, 345-363. R Core Team. (2017). R: A language and environment for statistical computing. R Found Stat Comput Vienna, Austria URL. Redmond, K. J., Berry, M., Pampanin, D. M., & Andersen, O. K. (2017). Valve gape behaviour of mussels (Mytilus edulis) exposed to dispersed crude oil as an environmental monitoring endpoint. Marine Pollution Bulletin, 117(1-2), 330- 339. Rezek, R. J., Lebreton, B., Roark, E. B., Palmer, T. A., & Pollack, J. B. (2017). How does a restored oyster reef develop? An assessment based on stable isotopes and community metrics. Marine Biology, 164(3), 54. Riascos, J. M., Avalos, C. M., Pacheco, A. S., & Heilmayer, O. (2012). Testing stress responses of the bivalve Protothaca thaca to El Niño-La Niña thermal conditions. Marine Biology Research, 8(7), 654-661. doi:10.1080/17451000.2011.653367 Richkus, W. A., & Menzie, C. A. (2013). Application of an Ecological Risk Assessment for Evaluation of Alternatives Considered for Restoration of Oysters in Chesapeake Bay: Background and Approach. Human and Ecological Risk Assessment: An International Journal, 19(5), 1172-1186. doi:10.1080/10807039.2013.767092 Ricker, W. E. (1975). Computation and interpretation of biological statistics of fish populations. Bull. Fish. Res. Bd. Can., 191, 1-382.

54

Riisgård, H. U., Lassen, J., & Kittner, C. (2006). Valve-gape response times in mussels (Mytilus edulis)—effects of laboratory preceding-feeding conditions and in situ tidally induced variation in phytoplankton biomass. Journal of Shellfish Research, 25(3), 901-912. Rinkevich, B. (2014). Rebuilding coral reefs: does active reef restoration lead to sustainable reefs? Current Opinion in Environmental Sustainability, 7, 28-36. doi:https://doi.org/10.1016/j.cosust.2013.11.018 Rodland, D. L., Schöne, B. R., Baier, S., Zhang, Z., Dreyer, W., & Page, N. A. (2008). Changes in gape frequency, siphon activity and thermal response in the freshwater bivalves Anodonta cygnea and Margaritifera falcata. Journal of Molluscan Studies, 75(1), 51-57. Rose, T. H., Tweedley, J. R., Warwick, R. M., & Potter, I. C. (2019). Zooplankton dynamics in a highly eutrophic microtidal estuary. Marine Pollution Bulletin, 142, 433-451. doi:https://doi.org/10.1016/j.marpolbul.2019.03.047 Rubio, A., Winberg, P., Kirkendale, L., & Warner, R. (2013). Ensuring that the Australian Oyster Industry adapts to a changing climate. Ruesink, J. L., Lenihan, H. S., Trimble, A. C., Heiman, K. W., Micheli, F., Byers, J. E., & Kay, M. C. (2005). Introduction of non-native oysters: ecosystem effects and restoration implications. Annu. Rev. Ecol. Evol. Syst., 36, 643-689. Rybovich, M., La Peyre, M. K., Hall, S. G., & La Peyre, J. F. (2016). Increased temperatures combined with lowered salinities differentially impact oyster size class growth and mortality. Journal of Shellfish Research, 35(1), 101- 114. Saville-Kent. (1893a). Albany oyters Mr Saville-Kent's report. The West Australia. Retrieved from http://nla.gov.au/nla.news-article3056333 Saville-Kent. (1893b). Oyster fisheries in the estuary of the Swan. The West Australian. Retrieved from http://nla.gov.au/nla.news-article3047161 Scyphers, S. B., Powers, S. P., Heck Jr, K. L., & Byron, D. (2011). Oyster reefs as natural breakwaters mitigate shoreline loss and facilitate fisheries. PloS one, 6(8), e22396. Shumway, S. (1977). Effect of salinity fluctuation on the osmotic pressure and Na+, Ca 2+ and Mg 2+ ion concentrations in the hemolymph of bivalve molluscs. Marine Biology, 41(2), 153-177. Smaal, A., Verbagen, J., Coosen, J., & Haas, H. (1986). Interaction between seston quantity and quality and benthic suspension feeders in the Oosterschelde, The Netherlands. Ophelia, 26(1), 385-399. Steckis, R. A., Potter, I. C., & Lenanton, R. C. J. (1995). The commercial fisheries in three south-western Australian estuaries exposed to different degrees of eutrophication. In: CRC Press U6. Taylor, A. M., Maher, W. A., & Ubrihien, R. P. (2017). Mortality, condition index and cellular responses of to combined salinity and temperature stress. Journal of Experimental Marine Biology and Ecology, 497, 172-179. doi:10.1016/j.jembe.2017.09.023 Therneau, T. (2015). A Package for Survival Analysis in S. version 2.38. Tolley, S. G., & Volety, A. K. (2005). The role of oysters in habitat use of oyster reefs by resident fishes and decapod crustaceans. Journal of Shellfish Research, 24(4), 1007-1013.

55

Tweedley, J. R., Hallett, C. S., & Beatty, S. J. (2017). Baseline survey of the fish fauna of a highly eutrophic estuary and evidence for its colonisation by Goldfish (Carassius auratus). International Aquatic Research, 9(3), 259-270. doi:10.1007/s40071-017-0174-1 Tweedley, J. R., Hallett, C. S., Warwick, R. M., Clarke, K. R., & Potter, I. C. (2016a). The hypoxia that developed in a microtidal estuary following an extreme storm produced dramatic changes in the benthos. Marine and Freshwater Research, 67(3), 327-341. doi:https://doi.org/10.1071/MF14216 Tweedley, J. R., Warwick, R. M., Clarke, K. R., & Potter, I. C. (2014). Family-level AMBI is valid for use in the north-eastern Atlantic but not for assessing the health of microtidal Australian estuaries. Estuarine, Coastal and Shelf Science, 141, 85-96. doi:https://doi.org/10.1016/j.ecss.2014.03.002 Tweedley, J. R., Warwick, R. M., & Potter, I. C. (2016b). The contrasting ecology of temperate macrotidal and microtidal estuaries. In Oceanography and Marine Biology (pp. 81-180): CRC Press. Tweedley, J. R., Warwick, R. M., Valesini, F. J., Platell, M. E., & Potter, I. C. (2012). The use of benthic macroinvertebrates to establish a benchmark for evaluating the environmental quality of microtidal, temperate southern hemisphere estuaries. Marine Pollution Bulletin, 64(6), 1210-1221. doi:10.1016/j.marpolbul.2012.03.006 Valesini, F. J., Hallett, C. S., Hipsey, M. R., Kilminster, K. L., Huang, P., & Henning, K. (2019). The Peel Harvey Estuary. In Coasts and Estuaries: The Future (pp. 103-120): Elsevier. van Keulen, M. (2002). Seagrass Transplantation in a High Energy Environment. Experiences from Success Bank, Western Australia. Paper presented at the Restoration Workshop for Gulf St Vincent 15–16 May 2001. Venture, A. (2016). Edible Oyster Aquaculture in the and Gascoyne Regions of Western Australia: A preliminary feasibility assessment. Retrieved from Verdelhos, T., Marques, J. C., & Anastácio, P. (2015). The impact of estuarine salinity changes on the bivalves Scrobicularia plana and Cerastoderma edule, illustrated by behavioral and mortality responses on a laboratory assay. Ecological Indicators, 52, 96-104. doi:10.1016/j.ecolind.2014.11.022 Verduin, J., Paling, E., & van Keulen, M. (2010). Seagrass rehabilitation; requisites to successful, long term, mitigation outcomes. Warton, D. I., & Hui, F. K. C. (2011). The arcsine is asinine: the analysis of proportions in ecology. Ecology, 92(1), 3-10. doi:10.1890/10-0340.1 Warwick, R. M., Tweedley, J. R., & Potter, I. C. (2018). Microtidal estuaries warrant special management measures that recognise their critical vulnerability to pollution and climate change. Marine Pollution Bulletin, 135, 41-46. doi:10.1016/j.marpolbul.2018.06.062 Wilcox, M., Kelly, S., & Jeffs, A. (2018). Ecological restoration of mussel beds onto soft‐ sediment using transplanted adults. Restoration Ecology, 26(3), 581-590. doi:10.1111/rec.12607 Wildsmith, M. D., Rose, T. H., Potter, I. C., Warwick, R. M., & Clarke, K. R. (2011). Benthic macroinvertebrates as indicators of environmental deterioration in a large microtidal estuary. Marine Pollution Bulletin, 62(3), 525-538. doi:10.1016/j.marpolbul.2010.11.031 Wildsmith, M. D., Rose, T. H., Potter, I. C., Warwick, R. M., Clarke, K. R., & Valesini, F. J. (2009). Changes in the benthic macroinvertebrate fauna of a large

56

microtidal estuary following extreme modifications aimed at reducing eutrophication. Marine Pollution Bulletin, 58(9), 1250-1262. doi:10.1016/j.marpolbul.2009.06.008

57