Feasibility of shellfish 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

iii

Abstract

With 85% of oyster 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 Australia 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-western Australia. 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 Saccostrea glomerata, 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

iv

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.

v

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 oysters ....................................................................................... 6

1.4 Oyster reef restoration .............................................................................................. 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

vi

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 Mussel 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.

1

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

2

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 seagrass 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 mussels (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 (Ostrea

3

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. Queensland 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 South Australia 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

4

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 Crassostrea

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, Ostrea edulis, which produced an

5

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;

6

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 animal subsequently reduces activities related to respiration,

7

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 aquaculture. 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

8

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 Rock Oyster, 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 New South Wales 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 Tasmania harbours a historic O. angasi reef (Gillies et al., 2018). This

9

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 Northern Territory 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 Pacific Oyster 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).

10

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

11

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

12

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.

13

Chapter 2: Materials and Methods

2.1. Peel-Harvey Estuary

The Peel-Harvey Estuary, located 80 km south of the City of Perth 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

14

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.

15

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

in Western Australia.

.

16

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).

17

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)

18

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].

19

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,

20

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

21

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.

22

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):

BCI=dry weight (g)

shell length (mm) × 100

23

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 ∆𝑀𝑗 = 𝑀𝑡𝑗 − �̂�𝑏, where ∆𝑀𝑗 is the change in BCI of the j’th oyster, 𝑀𝑡𝑗

is the BCI of the j’th oyster after treatment and �̂�𝑏 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.

24

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).

Fig. 2.5 Pearson correlation test between the proportion of time spent open at weekly intervals of

the trial.

0

0.2

0.4

0.6

0.8

1

0 1 2 3 4 5 6 7 8

Pro

po

rtio

n o

f tim

e s

pe

nt

op

en

Weeks of trial

25

Fig. 2.6 Flowchart illustrating the experimental process.

26

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

27

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.

28

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].

29

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.

30

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 10 th 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)

31

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).

32

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

33

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.

34

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.

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

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

35

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.

a)

b)

36

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.

37

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.

38

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.

39

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

40

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).

41

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.

42

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)

43

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 animals 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.

44

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 seagrasses, 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

45

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

46

(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 eastern oyster 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).

47

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.

48

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.

49

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