Assessing and mitigating environmental impacts of SWRO outfalls on key benthic marine organisms

Transcription

1 Assessing and mitigating environmental impacts of SWRO outfalls on key benthic marine organisms Julie Mondon, Marion Cambridge, Anne Brearley, Gary Kendrick May 2016 National Centre for Excellence in Desalination Australia Project #08699

2 Julie Mondon 1* Marion Cambridge 2 Anne Brearley 2 Gary Kendrick 2 1. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment, Deakin University, Warrnambool, Victoria Australia. Julie.mondon@deakin.edu.au 2. Oceans Institute, School of Plant Biology, University of Western Australia, Crawley, Western Australia. Page 1 215

3 ACKNOWLEDGMENTS AND PERSONNEL The Assessing and mitigating environmental impacts of SWRO outfalls on key benthic marine organisms project has been funded by the National Centre of Excellence in Desalination Australia, NCEDA Project code # Multiple institutions and individuals have contributed to the project; overall project design, field surveys, experimental exposure trials, data collection and analysis, and information synthesis. Deakin University: Dr Julie Mondon, Dr Amanda Bates, Dr Alecia Bellgrove, Stephen Wheeler, Shaun Davis. University of Western Australia: Dr Marion Cambridge, Prof. Gary Kendrick, Dr Anne Brearley, Dr John Statton, Andrea Zavala-Perez, Renae Hovey National Centre for Excellence in Desalination: Dr Misty Palmer Water Corporation: Laura Chidgzey Project leader: Dr Julie Mondon Report production: Dr Julie Mondon Original project design: Dr Julie Mondon, Dr Marion Cambridge, Prof. Gary Kendrick, Dr Amanda Bates, Dr Alecia Bellgrove. Field survey 2012: Dr Julie Mondon, Dr John Statton, Dr Marion Cambridge, Prof. Gary Kendrick, Stephen Wheeler, Andrea Zavala-Perez, Davide Abate Field survey 2013: Dr Anne Brearley, Andrea Zavala-Perez, Shaun Davis, Todd Bond, Damon Driessen, Sam Gustin-Craig, Kim Royce (Skipper Cross Country ), David Whitham (Ocky s Dive World), Davide Abate, Claire Ross Field survey statistical analyses: Dr Anne Brearley, Prof Gary Kendrick, Andrea Zavala-Perez Seagrass salinity and brine trials analyses: Dr Marion Cambridge, Dr John Stratton, Andrea Zavala-Perez, Renee Hovey, Greg Cawthray, Dr Julie Mondon Mussel salinity and brine trials analyses: Dr Julie Mondon, Shaun Davis, Elizabeth Cain Diffuser community survey: Dr Julie Mondon, Dr Anne Brearley, Freemantle Dive (Antony Old, Carly Bertolino, Ben Fazioli, diver Mark Chinkin) Diffuser mussel health condition evaluation: Dr Julie Mondon, Dr Anne Brearley International Expert Advisory team: Prof Daniel Schlenk (University of California, Riverside), Dr Juan Manuel Ruiz (Oceanography Centre of Murcia, Spain) Project logistics, industry liaison, data collection and imagery support: Dr Misty Lee-Palmer (NCEDA), Laura Chidgzey (Water Corporation), Tymen Brom (NCEDA) Blue mussels supply: Bunbury Mussel Farm, Bunbury WA; Advance Mussel Supplies, Port Arlington VIC Chemical analyses: Water Corporation WA, Water Quality Laboratory (Deakin University, Warrnambool) Page 2 215

4 Contents ACKNOWLEDGMENTS AND PERSONNEL... 2 Contents Project overview and scope... 6 Background (technical basis) Key recommendations for impact assessment of brine discharge Environmental Impact Multiple Lines of Evidence Assessment Approach Summary of methodological approach The Model desalination plant... 9 Field Impact Assessment Receiving Environment Impact Assessment Benthic habitat Field sampling considerations Gradient design Biological component Physico-chemical considerations Biomarker development Biomarkers to evaluate biological effect Critical threshold stress responses Field validation of biomarker stress responses Biomarker selection Physiological / behavioural biomarkers Histological biomarkers Utility of the biomarker approach Linking desalination waster brine discharge to biological effect References Field Assessment Invertebrate fauna in seagrass and reef pavement habitats in the vicinity of a desalination plant at Binningup, Geographe Bay southern Western Australia Introduction Methodology Results Discussion References Page 3 215

5 3. Biomarker development Seagrass Morphological and physiological effect of elevated seawater salinities and desalination-derived brine on key seagrass species Overview Pilot study Elevated seawater salinity tolerance trials Mature seagrass brine exposure Seedling brine exposure Summary of response to elevated seawater hypersalinity and brine References Biomarker development - Bivalves Morphological and physiological effect of elevated seawater salinities and desalination-derived brine on key bivalve species Overview Pilot investigation Hypersalinity stress response - Xenostrobus pulex Seawater hypersalinity stress response - Mytilus edulis sp Desalination brine stress response - Mytilus edulis sp Summary of findings References Field evaluation Community structure alteration Investigation of in situ biota associated with desalination brine diffuser discharge Introduction Methodology Results Riser and surrounding seabed community composition Discussion Summary Overview - Use of Kwinana diffuser benthic community as a proxy for Binningup References Field evaluation Bivalve health condition, Kwinana desalination diffuser outfall Introduction Methodology Results Summary and implications of findings References Recommendations Biomarkers of effect Page 4 215

6 7.2 Bivalves - Morphological and tissue-level assessment of health condition Bivalves - (Physiological - metabolic assessment of health condition) Bivalves Considerations for transplantation of mussels for monitoring Bivalves References Seagrass Biochemical, tissue-level and whole plant response to marine hypersalinity Seagrass - Whole organism level Seagrass - Tissue and cellular level Seagrass - Biochemical and biophysical level Seagrass - Developmental level Seagrass - References Field surveys - In situ evaluation Page 5 215

7 1. Project overview and scope Desalination brine has been identified to disturb both physico-chemical and ecological attributes of receiving marine environments [Winters et al. 1979, Maugin & Corsin 2005, Miri & Chouikhi 2005]. Environmental concerns relating to the impact of SWRO outfalls include the physical and chemical impacts of hypersaline discharges upon the receiving system. Approximately 40 50% of the total volume of seawater used by SWRO desalination plants results in a hypersaline (40 70 psu) waste (brine) which is discharged. High salt concentration is a potential stressor for benthic communities in receiving aquatic ecosystems [Bath et al. 2004]. The use and release of trace anti-foulants and anti-scalants to maintain plant infrastructure [Ketsetzi et al. 2008], and the potential concentration of existing contaminants into fluids during brine concentration generates other potential stressors to local marine life. Growing global dependence on desalination to generate freshwater resources (up to 25 million m 3 of desalinated water is produced daily around the world [Lattemann & Hopner 2008]), means that there is a need to increase the current paucity of knowledge relating to environmental impacts of waste brine discharge on aquatic organisms [Lattemann & Hopner 2008, Tulharam & Ilahee 2007]. Specifically, in relation to the Australia context at the time when this project was initiated, limited information on direct toxicity of desalination brine on coastal marine species was available, and to some extent contradictory [Roberts et al. 2010, RPS 2009]. At present no environmental guidelines based on species tolerances to desalination brine discharge into nearshore coastal habitats exist. In Western Australia, brine discharge into seagrass meadow habitats has elicited strong public concern. Seagrass meadows are highly productive communities, providing valuable ecological and socio-economic functions and services to estuarine and marine coastal ecosystems. These meadows support a diverse community of invertebrates, including the bivalves, and benthic finfish, the latter two often being utilized extensively by recreational fishers. Notably, almost one third of the world s seagrass species occur in southwestern Australia, and these seagrass communities play a major role in the marine ecology, with benthic primary production from seagrasses and algae supporting detrital food webs [Carruthers et al. 2007]. The population of Page 6 215

8 Western Australia lives in close proximity to the sea, and together with governmental and nongovernmental institutions, they share a genuine and serious concern for the marine environment. As a consequence, sea grass meadows are identified as priority ecosystems for conservation and protection [Zann 1995, 1996]. Evidence suggesting that seagrass meadows, and the macrofaunal community they support, may be vulnerable to changes in salinity regimes induced by hypersaline effluents from desalination plants is documented [Carruthers et al. 2007], but specific knowledge of the tolerance of Australian marine species to rapid alterations in salinity and temperature is limited. Furthermore, potential synergistic stress response that may occur due to hypersalinity is unknown. In providing a detailed understanding of key species biological response/s to brine, and the interaction of brine with in situ physico-chemical characteristics in the receiving environment, bioindicators and biomarkers of effect can be identified and which can be used by the desalination industry in planning and regulation of SWRO discharge into receiving waters. The Binningup desalination project concept arose in part from community concern relating to potential environmental impact from Western Australian desalination plants. It was envisaged that the Western Australian data could provide a baseline model for biological risk assessment of desalination discharge to coastal habitats across Australia, and serve to shape future mitigating strategies by filling essential knowledge gaps that could be transferred from one geographical region to another. Moreover, to gain the trust, engagement and ongoing support of local communities, irrespective of location, it is essential that detailed understanding of immediate and long-term marine ecosystem impacts can be effectively evaluated, and communicated to allay community concern. Background (technical basis) - Key recommendations for impact assessment of brine discharge The three key recommendations for effective environmental assessment of desalination waste brine discharge, proposed by the Environmental Impacts of Desalination in California working group [Kendrick et al. 2008], were considered to be directly relevant to the Australian situation: Page 7 215

9 1. there is a clear need to study the near-shore impacts of brine discharge on local species, 2. it is necessary to consider various levels of biological organisation when measuring such impacts, and 3. there may be interacting effects from hypersaline concentrations and other physicochemical stressors present in the brine responsible for impacts that need to be identified [Alper et al. 2007]. Based on the key recommendations, identification of a brine discharge, biologically significant change at the individual through to community level, and identification of causal relationships between stress response and brine exposure, provide a knowledge base to generate an assessment tool box applicable for brine discharge risk management. The tool box, using a multiple lines of evidence approach, is aimed to answer five broad key research questions with particular reference to Western Australian desalination outfalls: 1. the potential impact of brine on the benthic community in receiving waters 2. the spatial scale of impact 3. the potential temporal scale of impact 4. the specific effects of brine toxicity at hierarchical levels of biological organisation 5. the potential for concomitant increase or decrease in stress response due to interaction between brine and the physico-chemical characteristics of the receiving waters. - Environmental Impact Multiple Lines of Evidence Assessment Approach Environmental impact assessment of chemical stressors in aquatic ecosystems requires a multitiered, multiple lines of evidence approach [Alper et al. 2007]. Multiple lines of evidence based on the presence of anthropogenic-derived chemical/s and / or physical stressors in the receiving environment, evidence of exposure and /or assimilation of those stressors by organisms, and the presence of a deleterious biological response, enables the potential risk from such exposures to be confidently determined [Schlenk 2003]. - Summary of methodological approach Aquatic organisms will be investigated, both as integrators of stress effects and as sensitive response (early warning) indicators of environmental health. The methodological approach will utilise controlled in-situ field assessments and laboratory experiments. A field Impact Assessment will be used to identify bioindicators of response, ie. alteration at the population and community levels. Sentinel animal and plant species will be used to develop biomarkers Page 8 215

10 of stress response at the individual level. Biomarker development will involve a two-stage controlled laboratory exposure and field evaluation program. - The Model desalination plant Southern Seawater Desalination Plant (SSDP) at Binningup has been chosen as the model SWRO facility. This plant is preferred over the Perth Seawater Desalination Plant at Kwinana, Cockburn Sound. Binningup is more recent (September 2011), and representative of a direct discharge operation to open coastal waters. Additional anthropogenic impact from industrial activity in the region of the Binningup desalination plant is limited, relative to the Perth Seawater Desalination Plant at Cockburn Sound. - Background knowledge of SSDP Potential impacts on the marine environment associated with the brine discharge at Binningup have been identified and reported by the Water Corporation [Water Corporation]. Projected risk and severity of risk is based on knowledge from other desalination operations and similar likelihood of the same occurring at Binningup. At the time of the report s release, tests to evaluate and confirm the projected outcomes had not been conducted. The following information is based on the Report s evaluation of the following potential risks identified: - Reduction in ph: The discharge ph will range from 6 to 8, and is likely to convert to background levels rapidly. Increase in nutrients from processing chemicals (polyelectrolytes, biocides and acid detergents), causing eutrophication: Several chemical used in the process are potential sources of biologically available nitrogen and phosphorus. It was concluded that the Binningup plant is unlikely to increase nitrogen levels in marine waters [this has not been directly tested]. Increase in salinity affecting flora and fauna: Marine benthic invertebrates have generally been found to be tolerant of slight (up to 5 psu) increases in seawater [abeit over short term exposures]. The ANZECC/ARMCANZ (2000) National Water Quality Guidelines are not specific for desalination brine and recommend that biological effects using local reference data (mainly physical and chemical stressors) be used to identify salinity stress tolerances in preference to default trigger values. Page 9 215

11 Chemicals in brine affecting flora and fauna: Several chemicals are listed as potential stressors; Ferric Sulphate/Chloride, Polyelectrolyte, Ferric sulphate / ferric chloride, Antiscalants (active ingredient Nalco product PC 1020 phosphinocarboxlic acid), sodium, Hypochlorite, Sodium Metabisulphite, Biocide (Occtech OE-BIO-2000 broadspectrum 2,2- dibromo-3-nirilopropion amide (DBNPA)), Citric Acid or Caustic acid, and Metals. It is predicted that due to the dilution factor, comparability to naturally occurring materials, degradation processes to natural by-products, and neutralisation processes in seawater, most of these chemicals are not predicted to cause concern. The report did state that reverse osmosis desalination plants do measurably add metals to the brine, but the concentration is then diluted. Metals concentration in the intake seawater will approximately double in concentration before being discharged with diluent water in the brine stream. The dilution factor would reduce the level to slightly (estimate of 4%) above ambient concentration. This percent concentration is not high, however it was recommended that monitoring of the discharge metals concentrations be undertaken. - Lower dissolved oxygen (DO) levels affecting flora and fauna: There is not likely to be a significant change, beyond natural variation, in any ecological or biological indicators that are affected by poorly oxygenated water in deeper waters. However it was also noted that the solubility of oxygen decreases in seawater with increased temperature, and the SSDP brine will be up to 2ºC warmer and up to around 30 ppt saltier than seawater. Additionally, sodium bisulphite, used to remove residual chlorine, can also lower DO. Further, it has been estimated that oxygen depletion could decrease up to 2 to 2.5 mg/l lower in DO relative to the ambient seawater. Sustained DO levels at the seabed will further depend on stratification events, in which it has been estimated that conditions suitable for stratification could occur approx 20 25% of the time at Binningup (UWA research). It has been considered that dissolved oxygen could be reduced in zones below a density or thermally induced stratification, however the risk is predicted to be low. In addition to salinity and temperature, DO was monitored at the diffuser at regular intervals over the first year of production to enumerate actual levels to the receiving environment (Water Corporation). Page

12 Field Impact Assessment - Receiving Environment Impact Assessment Benthic habitat A comprehensive spatial short duration brine exposure vs prolonged duration brine exposure field evaluation will be conducted, premised on the traditional BACI (Before After Control Impact) design. The Binningup plant began 30% operational discharge in September It was not possible to survey impact of brine discharge using a strictly traditional BACI design as Before samples cannot be taken. However, the principles of advanced BACI designs [Benedetti-Cecchi 2001, Underwood 1991, 1992] allow for the effects of short-term exposure vs prolonged exposure to be assessed. - Field sampling considerations Field sampling considerations taken into account were the following: 1. a number of potentially confounding influences in the area (e.g. sewage effluent discharge, mineral sands mining, the Peel-Harvey estuarine outflow); 2. natural spatial and temporal variability in the distribution and abundance of marine organisms in the area; 3. the relationships between marine biota, substratum characteristics and depth; 4. reliability of weather conditions for boat-based SCUBA surveys; 5. timeframe of the project; and 6. budgetary constraints. The sampling design (Fig. 1) provides the strongest statistical power to detect effect/s from brine discharge on in situ marine biota. The design has been developed within the framework of the considerations outlined above, and the geographical position of the Southern Seawater Desalination Plant at Binningup. - Gradient design Gradient designs establish sites on a gradient of stressors which is equivalent to the traditional dose response model but based on field exposure. Sites are assessed by comparing the assemblage of organisms with increasing distance from the outfall, to an expectation derived from relatively undisturbed reference sites (furthest from the outfall). Identification of suitable reference sites is based on physico-chemical and geographic information. Nine geo-located sites varying in distance and direction (north and south) from the Binningup outlet (Fig. 2), comprising three habitats (high-relief reef, sand inundated limestone pavement Page

13 and mixed seagrass-non vegetated sand habitats) that incorporate the main substratum types in this region [Kendrick et al 2008], are sampled using replicated belt transect surveys, controlling for depth. - Biological component Dive teams are used to quantify the biomass of benthic plants (seagrasses, seaweeds), and identify and quantify numbers of sessile/sedentary animals (invertebrates, territorial finfish) along each transect. Primary focus on sessile/sedentary benthic species stems from benthic organisms being the most vulnerable to potential discharge impacts. Sessile and sedentary organisms have limited physical and /or behavioural capability to avoid adverse conditions generated from a plume discharge. Benthic species living in and around seagrass meadow habitats in close proximity to desalination outfalls are potentially exposed to hypersaline water depleted in oxygen, nutrients and food. Fig. 1 Sampling design for field-based impact assessment of impact of brine discharge on marine biota. Page

14 Google Map Fig. 2 The geospatial position of each sampling site is represented by the specific site colour identified in the sampling design represented in Fig Physico-chemical considerations In situ physico-chemical characteristics of the water column at the Binningup site is frequently monitored (WaterCorp). Semi-permanent temperature and salinity loggers were also deployment to assess temporal fluctuations in the brine-effluent plume and the spatial extent Page

15 of dispersion from the source during this investigation. Concurrent water column profiles were aso conducted at each site during annual benthic community surveys, recording standard water quality parameters (temperature, salinity, ph and dissolved oxygen) to assess a shap shot of fine-scale variability of the physico-chemical environment during actual surveys. Biomarker development Determination and comparison of species composition across a stressor gradient allows the use of species diversity as an ecotoxicological endpoint to detect stress response. However, ecological reaction at the population and community can be complex and variable. Confirmation of specific biological response to stressors is supported by independent test data. Identification of causal relationships specific to critical threshold stress response needs to be established using biomarkers of stress in key benthic species. Physico-chemical exposure, assessed concurrently with alteration in biological structure / function, will enable identification of specific biological stress response to desalination brine exposure. - Biomarkers to evaluate biological effect Biomarkers based on alteration in biological function, which range from the biochemical to whole organism level, offer remarkable utility for development as indicators of exposure effect [Galloway 2006]. Stress responses at the tertiary level involve systemic changes in which organisms may become incapable of adapting to stressors, leading to adverse impact on overall health (Harper &Wolf, 2009), often indicated by reduced growth and maintenance, suppressed disease resistance, or reduced reproductive capability. A suite of biomarkers is also necessary to counter the variability in response expected between biological organization levels and species [Schlenk 2003, Galloway 2006]. Biomarkers indicating stress response (ie. biomarkers of effect), have been chosen to evaluate the toxicity potential of brine [Depledge et al. 1993, Hyland 2006]. Endpoints closely related to survival, growth and reproduction offer the greatest relevance to predicting ecological impact [Depledge et al. 1993, Hinton & Lauren 1990]. Page

16 Stress response in key plant and animal species is assessed against a seawater salinity and desalination brine exposure (dose response) gradient. Potential interaction/s between salinity, temperature, and brine-derived contaminants, are of particular relevance to environmental risk assessment of desalination waste brine. - Critical threshold stress responses Physico-chemical characteristics of in situ coastal water provide the baseline data to which organisms are normally exposed. Deviation from optimal ranges of environmental parameters (eg. salinity) can elicit strong stress response indicators due to the homeostatic mechanisms of aquatic organisms being highly dependent on prevailing environmental conditions [Harper & Wolf 2009]. Evaluation of brine as a physico-chemical stressor are assessed by species-specific stress responses ie. the physiological and morphological alterations evident in response to adverse environmental stimuli. Sand and shallow reef species (from multiple hierarchical trophic levels), are used as representative near-shore community model organisms. Adaptive laboratory- based bioassays will determine species tolerances to desalination waste brine water, and specific response to the physico-chemical characteristics comprising desalination waste brine water. Once doseresponse thresholds have been established, dose-response trials will be applied to determine whether interactions between brine, and temperature variations, are likely to elicit antagonistic, additive or synergistic responses in the model organisms; interactions that are environmentally relevant given the presence of seasonal episodes of warmer water currents. - Field validation of biomarker stress responses Laboratory trials will provide the basis for evaluation of organism response to in situ desalination waste brine exposure. Evidence of causal impact on organisms in the receiving waters will be based on the presence of those alterations in biological function identified under laboratory conditions. The receiving environment at Binningup consists of shallow reef and sandy substrate. Species representative of these habitats are identified as model organisms: Page

17 Seagrass 1. Posidonia sp., a dominant seagrass species representative of a sedentary primary producer that forms extensive seagrass meadow habitat [Kendrick et al 2008]. Posidonia australis, has been cultured by the Oceans Institute for the last 3 years, and is the most widespread of the genus for the southern half of Australia. 2. Halophila ovalis, commonly known as dugong grass, typical of soft sand substrates, with a widespread marine distribution throughout the Indo-Pacific. Bivalves Mussel sp. inhabit shallow sub-tidal to intertidal rocky reef habitats [Edgar 1997], and have been selected in preference to Pinna bicolour (Razor Clam), which is found in sheltered sand habitat but not applicable to the open Binningup coastal region, and Glycymeris radians, 2-30 m depth, with a wide distribution from WA to NSW, but field collection of animals not guaranteed when required. Two mussel species are considered to be applicable for this study: 1. Mytilus sp., highly abundant in dense populations along coastal shorelines where suitable settlement habitat is present. They have a widespread southern Australian and global distribution, and are commercially available in the Binningup region. 2. Xenostrobus pulex, highly abundant in dense populations on rocky and other stable substrates, with a wide southern Australian and southern hemisphere distribution. It is also an intertidal species tolerant of prolonged submersion. Biomarker selection A suite of biomarkers is considered essential for a screening analysis of environmental impact [Bowen & Depledge 2006, Galloway 2006]. Generic and specific biomarkers applicable to different levels of biological organisation (ie. suitable for adaptation to bivalves and seagrass) have been identified as the basis for a far broader suite. Whilst the number of biomarkers and species that could be investigated is potentially very large, those that are applicable to the long term survival of key species, and are most likely to prove successful for this particular geographical site have been chosen. Page

18 - Physiological / behavioural biomarkers Generic biomarkers Biomarkers linked to maintenance of body mass, growth, reproduction, and survival, can be applied to organisms at all hierarchical levels. Bivalves exposed to salinities and brine are evaluated for: metabolic capacity (indicated by oxygen uptake) morphological and pathological health, feeding rate, and condition. Reduction in condition can transfer to behavioural changes such as: - reduced burrowing behaviour, which can then lead to increased vulnerability to predation. - reduced tolerance to extremes in environmental parameters, which is costly based on the elevation of energy required for supporting physiological coping mechanisms. - reduction in defence response (ie. slow valve closure response which can lead to increased predation) Reduction in condition can transfer to changes in energy reserves such as: - reduction in body mass (ie. thinner bivalves exhibit a low mass body index*), and - reduction in gonadal mass (ie. reduced reproductive capacity, identified using a gonadal index*). - reduction in locomotory activity (ie. reduced byssal thread production) *Indices for growth and survival are necessarily essential indicators of survivorship of the organism, and ultimately the potential longevity of the population. Establishing the combination of factors that lead to mortality indicates the absolute upper limit in environmental exposures that can be tolerated. Seagrass exposed to elevated salinities and brine are evaluated for: - survival of leaf-bearing shoots - differences in rates of leaf growth Leaf growth is sensitive to physiological stress. Leaf elongation rates (ie. measure of growth) can be depressed under stressful conditions. - Histological biomarkers Histological alteration (changes in cell, tissue and organ structure) is particularly relevant where the potential for synergistic or additive influence of multiple stressor exposures is present [Hinton & Lauren 1990], as in the case of desalination brine. Environmentally induced changes in tissue biochemistry and physiology can lead to structural alterations at the cellular and tissue level. These histological alterations represent the Page

19 cumulative effect of multiple stressor exposures, revealing sub-lethal to chronic effects not easily detected by other means of biological evaluation [Hyland 2006]. A further advantage of histological alteration as a biomarker lies in the intermediate position between identifying alterations occurring at the biochemical level, and at the whole individual level [Hinton & Lauren 1990]. Histological alteration most often represents a medium to longer-term, or even delayed contaminant exposure, where biochemical change (resulting in a physiological alteration) has been either severe or protracted enough to elicit a histological response, whereas a physiological alteration is often indicative of recent (within days) exposure (Schlenk pers. comm.). Histological approaches provide a complementary approach to assessing physiological threshold limits in in situ organisms. Changes in cell, tissue and organ structure are assessed in bivalve species; the pathologies (types of cell change) present however, could potentially differ. Identification of the overall health status of the organism includes both generic and specific alteration measures to generate health indices. Generic Generic progressive and regressive alterations in major organs assessed: - digestive gland, kidney, gill, mantle Specific Specific alterations linked to loss of function are targeted. Health indices Histopathological health is identified for each species based on the alteration/s present, and the severity of that alteration, ie. whether the alteration is reversible with change to improved environmental conditions [Bernet et al 1999]. Histological health condition specific to organisms, in this case mussel species, provide a baseline allowing for statistical comparison between sites across a spatial gradient, and across temporal scales. The scientific basis on which the multiple lines of evidence methodological approach is developed is briefly outlined in the points that follow. Application of biomarkers and bioindicators of environmental stress involves identifying a suite of relevant stress responses at each level of biological organisation in order to: - assess the effects of environmental stressors on organisms, Page

20 - predict future trends ie early warning indicators of change, - obtain insights into causal relationships ie mechanisms, between stress and effects at higher levels of biological organization, and - to provide a baseline with which to evaluate the effectiveness of remedial actions if necessary. Utility of the biomarker approach - Linking desalination waster brine discharge to biological effect. The biomarker field assessment approach advocated is designed to establish and quantify risk of exposure to key species in receiving waters, the tenet being that identification of adverse consequences for species occupying critical trophic positions will provide insight into the biological stability of the ecosystem as a whole [Depledge et al. 1993, Schlenk 2003, Galloway 2006]. Concurrent measures of in situ water chemistry, and biological stress response (biomarker, bioindicator and/or condition indices) can be used to identify potential links between hypersaline brine exposure in the environment, to biological response at the multi-organ, multispecies level. Concurrent measures allow: determination of brine concentrations (or complex contaminant mixtures within brine) most likely to elicit deleterious response at the whole organism and tissue levels identification of baseline biomarker response against which biological changes can be measured and monitored development of health condition indices using morphological and histological parameters, and the application of these health condition indices for future comparison within and between species of similar taxa. A multi-tiered biomarker field approach to evaluating the ecological impact of desalination waste brine discharge offers a significant development in environmental risk assessment for the wider desalination industry. Page

21 References Alper, H., C. Borrowman, and B. Haddad (2007) Evaluating Environmental Impacts of Desalination in California. Center for Integrated Water Research 27 July 2007 (Oct. 2010). Bath, A., B. Shackleton, and C. Botica (2004) Development of temperature criteria for marine discharge from a large industrial seawater supplies project in Western Australia. Water SA 30(5): Benedetti-Cecchi L (2001) Beyond Baci: Optimization of environmental sampling designs through monitoring and simulation. Ecological Applications 11: Bernet D, Schmidt H, Meier W, Burkhardt-Holm P, Wahli T (1999) Histopathology in fish: proposal for a protocol to assess aquatic pollution. Journal of Fish Diseases 22: Bowen, R.E and M.H. Depledge (2006) Rapid Assessment of Marine Pollution (RAMP). Marine Pollution Bulletin 53: Carruthers, T. J. B., W. C. Dennison, Kendrick, G. A., Waycott, M., Walker, D.I. Cambridge, M. L. (2007). Seagrasses of south-west Australia: A conceptual synthesis of the world s most diverse and extensive seagrass meadows. Journal of Experimental Marine Biology and Ecology 350: Deng X, Rempel MA, Armstrong J, Schlenk, D (2007) Seasonal evaluation of reproductive status and exposure to environmental estrogens in Hornyhead turbot at the Municipal Wastewater Outfall of Orange County, CA. Environmental Toxicology 22: Depledge MH, Amaral-Mendes JJ, Daniel B, Halbrook RS, Kloepper-Sams, Moore MN, Peakall DB (1993) The conceptual basis of the biomarker approach. In DB Peakall, Shugart LR Biomarkers. Springer, Berlin. Pp Edgar, G. J. Australian Marine Life. New Holland Publishers pp.560 Galloway TS (2006) Biomarkers in environmental and human health risk assessment. Marine Pollution Bulletin 53: Harper CL, Wolf J Morphologic effects of the stress response in fish. ILAR J 50: Hinton DE, Lauren DL (1990) Integrative histopathological approaches to detecting effects of environmental stressors on fishes. American Fisheries Society Symposium 8: Hyland K (2006) Biological effects in the management of chemicals in the marine environment. Marine Pollution Bulletin 53: Kendrick et al Characterising the marine benthic habitats of the proposed Southern Seawater Desalination Plant (SSDP) site: interpretation from underwater towed video & map interpolation. Report to KBR Australia. Page

22 Ketsetzi, A., A. Stathoulopoulou, and K. Demadis (2008) Being green in chemical water treatment technologies: issues, challenges and development. Desalination 223: Lattemann, S., and T. Hӧpner (2008) Environmental impact and impact assessment of seawater desalination. Desalination 220: Maugin, G., and P. Corsin (2005) Concentrate and other waste disposals from SWRO plants: characterization and reduction of their environmental impact. Desalination 182: Miri, R., and A. Chouikhi (2005) Ecotoxicological marine impacts from seawater desalination plants. Desalination 182: Roberts, D.A., E.L. Johnston, and N.A. Knott (2010) Impacts of desalination plant discharges on the marine environment: A critical review of published studies. Water Research 44(18): RPS (2009) Effects of a desalination plant discharge pm the marine environment of Barrow Island. Chevron Australasia Pty Ltd. August 2009 No. N Schlenk D (2003) The relationships of biochemical endpoints to histopathology and population metrics in feral flatfish species collected near the municipal outfall of Orange County, CA. Environmental Toxicology and Chemistry 22: Tulharam, G., and M. Ilahee (2007) Environmental concerns of desalinating seawater using reverse osmosis. Journal of Environmental Monitoring 9: Underwood AJ (1991) Beyond BACI: Experimental designs for detecting human environmental impacts on temporal variations in natural populations. Australian Journal of Marine and Freshwater Research 42: Underwood AJ (1992) Beyond BACI: the detection of environmental impacts on populations in the real, but variable, world. Journal of Experimental Marine Biology and Ecology 161: Water Corporation Factors_Operational_Impacts.pdfBunningup Winters, H., I. Isquith, and R. Bakish (1979) Influence of desalination effluents on marine ecosystems. Desalination 30: Zann, LP (Ed.), 1995 Our Sea, our Future. Major Findings of the State of the Marine Environment Report for Australia. Dep. Environ. Sp. Territories, Canberra, pp Zann, L. P. (1996). "The State of the Marine Environment Report for Australia (SOMER): process, findings and perspectives." Ocean & Coastal Management 33(1-3): Page

23 2. Field Assessment Invertebrate fauna in seagrass and reef pavement habitats in the vicinity of a desalination plant at Binningup, Geographe Bay southern Western Australia Anne Brearley 1, Andrea Zavala-Perez 1., John Statton 1, Julie Mondon 2., Shaun Davis 2, Alecia Bellgrove 2., Marion Cambridge 1., Gary Kendrick 1 1. Oceans Institute and School of Plant Biology, The University of Western Australia, Crawley, Western Australia 2. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment Deakin University Warrnambool, Victoria Australia Key questions 1. To what extent are physicochemical characteristics of receiving waters altered by desalination waste brine discharge? 2. To what extent does near shore biota differ in community composition with increasing distance from the desalination waste brine discharge outfall? 3. If significant differences are present, what species or groups of species are responsible for those differences? 4. Whether there has been a temporal change in physicochemical characteristics. 5. Whether there has been a temporal change in community composition with increasing distance from the desalination waste brine discharge outfall. 6. If significant differences are present, what species or groups of species are responsible for those differences? 2.1 Introduction The proposal to establish a desalination plant at Binningup was the impetus for this study. Concerns expressed in 2011 that effluent from the plant would affect seagrass communities and the biota associated with them also highlighted the lack of information about biota in the area. No benthic surveys had been conducted in the area north of Bunbury, and there was lack of data on the seagrass species Posidonia angustifolia, the most abundant species in the northern sector of the bay near Binningup, and no published surveys of the biota. Earlier studies in Geographe Bay, reviewed in McMahon et al. (1997) and Barnes et al. (2008) were carried out south-west of study area, where the seagrasses consist of Posidonia sinuosa and Amphibolis antarctica communities. The earliest detailed study of underwater habitats close to Bunbury consisted of underwater habitat mapping at the Dalyellup south of Bunbury in 1999 in preparation for construction of a waste-water pipeline from the Bunbury WWTP, Page

24 with some 52 km of video transects filmed over a coarse and finer scale grid (Cambridge and Kendrick 2000). This site lies south of Bunbury, approx. 30 km from the Binningup outfall and consisted of a complex mosaic of patches of seagrass and reef pavement, with occasional higher (0.5-1 m) reef patches. Seagrass biomass samples were collected but this study did not include measurements of seagrass density. Previous surveys (Cambridge & Kendrick 2000 Kendrick et al. 2008, Barnes et al. 2007) in this largely unmapped area, had reported that the major habitats consisted of sand with the seagrass Posidonia angustifolia, and small outcrops of pavement reef. Observations of isolated clumps of algae attached to firm substrate in otherwise sandy areas were considered evidence that pavement reefs underlay areas of sand which was regularly smothered by sand, and the burial or exposure of the underlying pavements was indicative of high sand movement along this exposed coast (Keesing and Heine 2006). Furthermore the smothering of the reef habitats was considered likely to influence the types and longevity of some of the biota. A number of factors historically and currently affect the coast in the Bunbury area and include, outflow of nutrient rich water from the Leschenault Estuary, construction of the ocean channel The Cut, and development and maintenance of the port that includes dredging that disturbs the seabed and increases turbidity. Effluent from the Laporte Titanium plant was discharged on to the ocean beach north of Bunbury from In the process of refining ilmenite to titanium dioxide, an acidic effluent (ph ) was produced with large amounts of iron in solution, together with a suspension of fine mud containing other ions including heavy metal (Brearley 2005). In March 2012 areas north of the port of Bunbury were affected by turbidity from a dredge spoil ground (located about 1-2 km offshore in 6-11 m depth). Dredging is an annual activity at the port, (report by SKM on dredge spoil dumping from Bunbury Port). Suspended fine sediment caused turbidity up to km north of the spoil grounds. This is likely to have an impact on biota, including around the Binningup outfall. Spoil dumping has been carried out since 1976, and although the long-term management report (SKM 2011) reported that there was no seagrass in the area, it is not clear if there was ever a detailed survey before Video mapping of benthic habitats in the vicinity of the proposed desalination outfall for SSDP at Binningup (Kendrick et al. 2008, Oceanica 2008) were also hampered by turbidity. The first aim of habitat and faunal survey was to validate and extend the earlier mapping surveys in the Binningup area and south of Bunbury and locate seagrass and pavement areas Page

25 (sites) to the north and south of the Desalination outfall area. The second aim was to record the abundance and diversity of the larger, predominantly sessile macrofauna in adjacent seagrass and pavement reef habitats that could be impacted by effluent from the desalination plant. The survey was designed to sample 4 sites south of the outfall (Sites 1-4), the outfall (Site 5) and 4 sites to the north of the outfall (Sites 6-9) covering an area 70 km in length (Fig. 1). The third and fourth aims were to assess which species could be likely to accumulate constituents of the effluent, and species suitable for experimental growth, and mortality investigations using brine effluent that might be used as bioindicators of environmental effects. The report details the 2012 survey that covers 70 km of coast north and south of the Binningup Desalination Plant and Outfall. This is the first report of the seagrass and pavement habitats and invertebrate communities in this region. The patterns revealed by the 2012 survey, which showed a dissimilarity of topography and faunal composition at the two sites south of Bunbury City and Harbour and The Cut, entrance to Leschenault Estuary, led in 2013 to a redefining of the area of potential influence. The sampling area in 2013 was refocused on a 13 km area north of Bunbury, centered on the SSDP Pipeline allowing an increase in the number of sites (Fig. 1). It should be noted that these surveys were conducted after the SSDP pipeline construction and plant testing phases that had included excavation of the seabed, installation of the pipeline and mooring pylons and preliminary discharge of effluent. 2.2 Methodology Much of the coastline north and south of the Binningup Desalination Plant has not been mapped systematically. Although the major habitats of seagrass, which varies in shoot density, pavement reef and sand have been defined, and the regular movement of sand and smothering of seagrass and rocky areas noted, specific locations for systematic sampling have been documented. The sampling design was therefore based on a gradient analysis of seagrass and pavement reef communities at sites of varying distance from the Plant. Site selection was based on depth soundings and towed videos conducted in March and April 2012 along a 70 km area to the north and south of the Plant within the depth contour approximately 1 km off shore (Table1, Fig. 1). Sites chosen where based on the mapping and the close proximity of the two major habitats seagrass P. angustifolia and pavement reef. In 2012 no pavement reef was located at Site 7 Myalup and Site 8 Lake Preston South. Page

26 Table I. Sites and locations Binningup, NCED Survey 2012 and Site Location UTM Latitude Longitude 1 Peppermint Grove 50 H E N S E 2 Dalyellup 50 H E N S E 3 Buffalo Rd 50 H E N S E 4 Outfall South 50 H E N S E 5 Outfall 50 H E N S E 6 Outfall North 50 H E N S E 7 Myalup 50 H E N S E 8 Preston Lake South 50 H E N S E Fig. 1 Study Sites from Site 1 south to Site 7. Corresponding to original sampling design. Site 1 Peppermint Beach, Site 2 Dalyellup, Site 3 Buffalo Rd, Site 4 Outfall South (South Outlet), Site 5 Outfall, Site 6 North Outlet, Site 7 Myalup and Site 8 Preston Lake South. Page

27 In the south, Sites 1 Peppermint Beach and Site 2 Dalyellup, 45 km and 30 km respectively from the outfall, and Site 7 Myalup and Site 8 Lake Preston South, 5 km and 13 km north of the outfall, were selected to represent similar habitats at a greater distance from the outfall. Site 3 Buffalo Beach (8 km) and Site 7 Myalup (7km) were defined as local reference sites, less than 10 km from the outfall. Site 9 a further 8 km from Site 7, was only sampled in 2012 due to access difficulties, and inclement weather (Fig. 2 & 3). Site 5 Outfall was located immediately offshore from the SSDP desalination plant but some distance from the discharge opening that lies within a restricted area. Site 4 was approximately 900 m south of Site 5, and Site 6 approximately 300 m north. Fig. 2 Original sampling design for Sites span approximately 70 km of coastline. Seagrass and limestone pavement reef were sampled at each site if present. Page

28 Fig. 3 Sampling design Focused on the Outfall, North and South, including Intermediate level Sites 3 and 7. Physical Environment Winds and swells in the Geographe Bay area Winds and swells that influence the operational diving survey, turbidity and mixing of the water column were examined using date obtained from the Bureau of Meteorology (BOM). Bunbury hourly wind records 5 March 2011 to 8 March 2013 were obtained from BOM. Data were in km/hour and converted to Knots. Water quality - temperature, salinity and dissolved oxygen 2013 In 2012 and 2013 temperature, salinity and dissolved oxygen were recorded at each site at the time of invertebrate sampling. Due to equipment failures, the 2012 records are not presented. Temperature loggers were installed in March and April 2012 at Sites 3, 4, 5, 6 and 7. Two loggers were retrieved in May 2013 at Sites 4, 50 m from Outfall, and Site 7 Myalup, 5 Km north of the Outfall and had recorded temperature every ten minutes until March At each site during the 2013 invertebrate community surveys, temperature, salinity and dissolved oxygen were measured throughout the water column to approximately 0.5 m from the substratum. Temperature and Salinity were recorded at 0.25 m intervals with an YSI Castaway CTD. Three temperature and salinity profiles were taken at each site and the average presented. At site 4 only two profiles were taken on the 5 th of March. Dissolved Oxygen was recorded at 1 m Page

29 intervals using an YSI 85 dissolved oxygen probe. The YSI oxygen probe was calibrated at the beginning of each profile and the membrane was replaced prior to the fieldtrip. On each site visit one dissolved oxygen profile was taken. As measurements were tied to the attempted diving surveys, the number of profiles varied with each site ( Table 2) every time the site was visited for diving surveys. On the 8 th March 2013, the last day of the invertebrate survey, CTD profiles of temperature and salinity were also made in the vicinity of Site 5 outside the Outfall restricted area on the 8 th of March. The survey commenced on the 5 th March 2013 following a period of moderate westerly winds and swells, conditions usually associated with turbulence and mixing of the water column. Over the following 4 days, wind direction changed to south south-easterly, the wind strength and the swell declined. On the 7 th March the winds were slight, swell was low and conditions calm. On the 7 th and 8 th March conditions in the morning were calm, but in the afternoon the wind direction changed to south-west and wind speed and the swell increased. At the time of the surveys, unbeknown at the time, brine discharge was suspended on 6 March 2013 and recommenced on 12 March Due to equipment failures the 2012 records are not presented. Five temperature loggers were installed in May 2012 at Sites 3, 4, 5, 6 and 7. Loggers were retrieved from Sites 4 and 7 in May 2012 and recorded temperature every 10 minutes until March Table 2 Dates, salinity, temperature and dissolved oxygen profiles were recorded at each site during the 2013 survey. Date Site 3 Site 4 Site 5 Site 6 Site 7 5/3/ /3/ /3/ /3/ Records of water quality including daily, weekly, monthly quarterly, weekly measurements were obtained from Water Corporation and compared to the sampling conducted during the 3-5 days of biotic communities and incorporated into the assessment of temperature, salinity and oxygen in the outfall area. Brine discharge was reduced on 6 th March and did not occur during the remainder of the field program (5-8 th March). Brine discharged recommenced on the 12 March. Page

30 Seagrass sampling The benthic survey for seagrass was carried out together with the faunal survey over 2 sampling phases in March and April 2012 and March 2013 at sites along the Geographe Bay coast. Shoot density and biomass quadrat sampling Seagrass, Posidonia angustifolia, was harvested by SCUBA divers for estimates of shoot density and above-ground biomass. All shoots with rhizomes attached were removed from one 20 x 20 cm quadrat (an area of 0.04 m 2 ) for each transect to give a total of 6 samples per site. Shoots were placed in calico bags and frozen until analysis. Average shoot density per site was estimated by counting the number of shoots from each quadrat, and then scaling up to 1 m 2. Leaf mass was measured by cutting the leaves from the leaf bases and then drying for 48 hours at 60 o C. Leaf dry weights were also scaled up to 1 m 2 and the 6 samples per site were then averaged. Leaf area per shoot was measured for 10 complete shoots harvested from the 6 replicate 20 x 20 cm quadrats. The length and width of leaves on each shoot were measured starting from the outer, oldest leaf (a), which enclosed successively younger leaves, (Fig. 4). The areas of individual leaves on each shoot were summed to calculate the leaf area per shoot. Leaf area index (LAI: m 2 leaf m -2 ground area) was calculated by multiplying the mean leaf area per shoot and the mean number of shoots per m 2. Oldest leaf (a) Mid leaf (b) Youngest leaf (c) Fig. 4 Arrangement of leaves on seagrass Posidonia plant. Longest, outer leaf (a) enclosing successively younger leaves (b, c). The oldest leaf is replaced every 3-6 months by a younger, shorter leaf (b). Page

31 Statistical analyses All values are given as means and standard errors (SE). Data were tested for normality and homogeneity of variances with Bartlett test and log-transformed. Differences in the variables of leaf biomass, leaf area and shoot density for the years 2012 and 2013 were tested for significance using a 1-way or 2-way ANOVA. If homogeneity of variances was not achieved with data transformation, then Permutational analysis of variance (PERMANOVA) test were carried out on a Euclidean distance resemblance matrix, constructed from the transformed data, using the PRIMER-E software package (Clarke et al. 2006). When a significant difference among variables was detected, pairwise test were carried to determine where the significant differences occurred among variables. Invertebrate sampling In 2012, at each site, amongst seagrass and on pavement (when present), six transect ropes five metres long, covering an area 30 m 2, were placed on the bottom. Diver operated videos were recorded along the length of the tape. The type and number of biota (Class level), within 0.5 m either side of the transect rope were recorded in situ on slates. Corresponding photographs were taken of the various species. A few voucher samples were collected for verification of diver records. Cryptic species were not sampled. Analysis of the videos, photographs of biota and the diver records indicated that although videos provided a general view of the areas, they did not effectively document the fauna due to the turbidity from the strong wave surge. The photographic and written records were more reliable but identification of the taxa was not ideal. In the April 2012 and March 2013, the diver swum videos were replaced systematic photographs of quadrats at each site. A one metre quadrat was placed at each metre point along each transects. The one metre quadrat or each 0.25 m 2 of the quadrat, dependent on visibility, was photographed to record habitat characteristics, dominant or unusual biota, and the number of major groups recorded on the sheets. We do not consider the modification of the sampling procedure compromised data collection for the remainder of the surveys. This was beneficial in terms of dive time and laboratory efficiency. The 2013 survey aim was to resample the seagrass and pavement at sites in the vicinity of the SSDP outfall and to increase replication within the area of the outfall (Sites 4, 5, and 6). Thus three sites sampled in 2012; Site 1 Peppermint Beach, and Site 2 Dalyellup and Site 8 Preston Lake South were not included in Page

32 Considerable movement of sand during large storms in late 2012 had however smothered much of the area and additional sites were only achieved in seagrass at Site 4a, and at Site 6a in an area of seagrass and fragmented pavement. These sites were approximately 50 m from the 2012 sites. Sand smothering also affected reference Site 3 Buffalo Road. The pavement and associated invertebrate community sampled in 2012 was not found and only one small area of seagrass was located, despite a wider search by divers who only encountered deep (1-20 cm) ripple marked sand. Seagrass and invertebrates could not be located at Site 5 nearest to the outfall due to the almost zero visibility. The already low visibility was further restricted when divers collected sediment. The fine sediment, black in colour with a strong odour was categorized as anoxic. Analysis This report documents the abundance of major taxa (mean ± SE) at each site. Each record has been attributed to particular taxa at species level or in the case of ascidians and sponges morphospecies. For the Ascidiacea species were based on shape (tubular, massive, branching, encrusting), colour, surface features (depressions, roughness, spines, reticulate ridges) and some common species identified with the aid of books by Gowlett-Holmes 2008, Edgar 1997, Shepherd and Davies1997, and the papers of Kott 1985, 1990 and 1992 and Porifera (sponge) species were defined by the presence and size of oscules, the internal matrix of the test (separate cortex and medulla), and consistency of the main body (soft, slimy, density and compaction). The defining internal skeleton, the spicules (megascleres and microscleres) were not examined. Background information was obtained from Goudie et al. 2013, Gowlett- Holmes 2008, Shepherd and Davies 1997 and Edgar Although identification of sponges is complex, recent studies using both morphological and species data indicate that differences between sponge communities using morphological diversity rather than species diversity may be used as a qualitative estimate of species diversity (Bell & Barnes 2001, Bell 2007, Bell et al. 2006). The effects of the physical environment (wave action, depth) to sponge morphology and size are acknowledged. Within a species, individuals exposed to high water movement can have a low profile and may be termed encrusting, whereas the same species in sheltered or deeper waters might have an upright form. Similarly shape is not consistent across genera. Despite these issues however, the exposure to high swells, and depth were similar at these sites with the exception of the most distant and southern site at Peppermint Beach. Page

33 Comparison of the community variability between and within sites was examined. The data matrix of site, habitats (seagrass and pavement) and species was examined using the PRIMER-E software package (Clarke et al. 2006) and the PERMANOVA + add on for PRIMER (Anderson et al. 2008). Transformed data were used to create a resemblance matrix based on the Bray-Curtis dissimilarity index. Analyses consisted in PERMANOVA and Canonical Analysis of Principal Coordinates (CAP). CAP was performed in order to characterise group differences found in the multivariate space. The Cross Validation gave the possible prediction of the distinctiveness of samples from each group. 2.3 Results Habitats and site Characteristics The 70 km survey area from the southern Site 1 Peppermint Beach to Site 8 Lake Preston South spanned more than 70 km of coastline (Fig. 1), a mosaic of sand and limestone reef pavement. A decreasing reef height from south to north and more extensive areas of sand in the north revealed a gradient in pavement and seagrass and the increased likelihood of sand smothering of the pavement communities. At some locations, seagrass predominately P. angustifolia grows in areas of sand either as small clumps in wave energy is high, the sand is mobile and lower areas of reef characteristically include small and large areas of sand. Similarly seagrass habitats included small sections of loose rubble and disturbance by divers of the sand veneer exposed the underlying pavement particularly in the sites nearest to the desalination plant. Episodic exposure and burial by the mobile sand of the hard substrate accounted for the presence of sessile taxa in unexpected habitats and could be defined as incidental occurrences or opportunistic settlement when conditions are favourable. The defined habitats (pavement and seagrass) of this survey could therefore be regarded as an over simplification and the biota found at the sites could be regarded as a gradient. Viewed together, the taxa recorded among seagrass and on pavement at each of the sites provide a good view of the species occurrence and abundance in the area (Fig. ). However at some sites areas of pavement were very small or not located and not sampled (see designs Figures 2 and 3). Page

34 Fig. 5 Habitats and characteristic biota surveyed (a-d) Site 1 Peppermint Beach showing seagrass, reef of higher relief with algae and corals. (e) Site 2 Dalyellup pavement reef of intermediate height, algal cover and large sponge typical of the area. (f) Pavement Site 3 Buffalo Rd, pavement overlain in part by sand and foliose algae attached to underlying pavement. (g) Pavement Site 5 Outfall (March 2012). (h) Seagrass Site 5 Outfall (March 2012). (i) Site 5 (March 2013). (j) Site 5 sediment (March 2013). Page

35 Physical Environment Wind and swells in the Geographe Bay area The most common winds were from the south and east. Strongly winds were predominantly from the west (Fig. 6). The lowest swells occur in the periods December to March. The plot of swell direction and strength illustrates the dominance of swells from the South West with a yearly average of about 2.5 m (Fig. 7a). Larger swells have a westerly trend (Fig 7b). (a) (b) Fig. 6 (a) Monthly average wind speed (± SD) from 5 March 2011 to 8 March Data obtained from BOM. Each month represents 750 wind measurements, standard deviations error term. (b) Proportional occurrence of hourly wind direction and strength from 5 March 2011 to 8 March 2013 at Bunbury, WA. Plotted from daily median swell forecast heights with standard deviation error term. Data from BOM. Page

36 (a) (b) Fig. 7 (a) Monthly average swell height forecast (± SD) from 5 September 2012 to 8 September (b) Proportional occurrence of swell direction and strength from 5 September 2012 to 8 September 2013 Between Dawesville and Cape Naturaliste, WA. Seasonal and annual changes in temperature Installed temperature loggers indicated a similar temperature pattern at Site 4 near the outfall and Site 7 the reference site 5 km to the north but ceased recording on 8th January The data points included winter and summer temperatures reaching a low of 15 C in August and maximum of 25 C in January (Fig. ). Page

37 (a) (b) Fig. 8 Temperature loggers installed Site 4 Outfall North 30th May and Site 7 Myalup (5 km to north of Outfall) 20thMay Loggers recorded to 8th January Five loggers were installed but only two were retrieved in March Brine discharge flow rates, physical properties and constituents Discharge rate, based on an estimation of 9,500 to 11,000 m 3 /hour, steadily increased throughout 2012 and beginning of 2013 (Fig. ). Between the two field surveys discharge was ceased for one extended period of approximately 20 days in August and September 2012 and two other occasions for approximately one day. Page

38 Fig. 9 Rate of discharge (m3/hr) of whole desalination brine effluent into marine receiving waters between 26/12/2011 and 30/04/2013. Period of sampling survey 2013 indicated. Mean salinity of brine effluent prior to discharge was approx psu (Fig. a). Properties of the brine effluent at discharge are presented in Table 3. At the discharge point the brine effluent is well oxygenated and ph similar to that of standard seawater. Startup conditions correlated with abnormally high ph and turbidity measurements on 18/09/2012, which declined to approximate mean values the next day. Several other high turbidity readings were also recorded, however no explanation of cause is noted in the data. Temperature of the effluent (Fig. b) followed a similar trend to the in situ temperature loggers at Site 4 and Site 7. Summer effluent temperature was approximately 25 C and winter effluent temperature slightly elevated at approximately 18 C. Two unexplained extremely high temperature readings were recorded on 22 December 2011 and 19 June 2012 at 101 C and 201 C respectively, were presumed to be errors in the data and excluded. Mean temperature of the effluent between 22 November 2011 and 25 September 2012 (removing two large outliers) was ± 0.22 C. Page

39 (a) (b) Fig. 10 (a) Salinity of whole brine effluent prior to marine outfall discharge from 22/11/2011 to 26/09/2012. (b) Temperature of whole brine effluent prior to marine outfall discharge from 22/11/2011 to 26/09/2012. Page

40 Metals and Nutrients analysis has been infrequent and variable sampling times occurred for different compounds between 22 November 2011 and 25 September Elevations in Boron, Bromine and Strontium appear consistent with concentration of standard seawater. Magnesium, sulphates, hardness are also concentrated but don t appear disproportionally inflated (Table ). Cr,Cu,Ni,Hg,As,Ag,Pb,Cd,Se have been below detection levels but only sampled on several occasions and detection limits may not be sensitive enough to pick up disproportional concentrations. For instance, standard seawater copper concentrations is approximately 0.001mg/l would require a 40 times increase to reach detection limit of 0.04 mg/l. Aluminium and Molybdenum were detected at concentrations of 0.59 and 0.05 respectively on one occasion otherwise below detection limit. Aluminium and Molybdenum are usually about and 0.01mg/L in seawater respectively. Residual Cl2 was measured on 87 occasions from 22 November 2011 to 25 September 2012 and was always recorded as 0.00 mg/l. Table 3 Whole brine effluent characteristics prior to marine discharge from 22/11/2011 to 26/09/2012. Parameter Mean ± SE Lowest Record Highest Record ANZECC Trigger Value+ Salinity g/l Not available ph * Dissolved Oxygen mg/l * 8.17^ Turbidity NTU * Not available +Table Default trigger values for physical and chemical stressors for south-west Australia (Australian and New Zealand Guidelines for Fresh ad Marine Water Quality Volume 1 October 2000). SSWA are to develop their own site specific trigger values based on over 12 months baseline monitoring. ^ Trigger value 90% saturation, conversion based on 20 C, altitude 0 m. *Rare extremely high value Page

41 Table 4. Concentration of additional parameters measured in desalination brine effluent prior to discharge. Parameter Number of samples Mean ± SE Boron mg/l Bromine mg/l Total Iron mg/l Strontium mg/l Total Suspended Solids mg/l Total Dissolved Solids mg/l Total Phosphorus µg/l Total Nitrogen µg/l Total Organic Carbon mg/l Chromium mg/l 3 <0.01 NA Copper mg/l 3 <0.04 NA Molybdenum mg/l 3 <0.01 but detected at 0.05 once on 23/4/12 Nickel mg/l 3 <0.04 NA Mercury mg/l 3 < NA Arsenic mg/l 3 <0.04 NA Silver mg/l 3 <0.04 NA Aluminium mg/l 3 <0.16 but detected at 0.59 on 19/01/2012 Lead mg/l 3 <0.04 NA Cadmium mg/l 2 <0.04 NA Magnesium mg/l Selenium mg/l 2 <0.06 NA Manganese mg/l NA Silica oxide µg/l NA Sulphate mg/l NA Nitrite mg/l 1 <0.002 NA Nitrate mg/l 1 <0.002 NA Ammonia µg/l NA Harness mg/l NA Temperature, salinity and oxygen profiles at five survey sites 2013 The survey data in March 2013 provides information on conditions at the individual sites on a short temporal scale of days during calm conditions that can influence mixing and stratification. Temperature Temperatures recorded at the sampling sites were generally slightly warmer on 5th March than subsequent days ( Page

42 Fig. ) when the sea was calmer with wind and swell easterly winds. Surface water was slightly warmer and there were slight increases in temperature occurring near the substratum at sites 4, 5 and 6 (closer to outfall). Page

43 Fig. 11 Temperature profiles at Binningup Sites 3-7. Three repeated profiles recorded at each site. Site 5 orange profile represents a second series of measurements taken later on 8th March Salinity Salinity at sites around outlet (Sites 4, 5, 6) all showed higher salinity on 5 March 2013 than subsequent days (Fig. 2). Site 5 Outfall, closest to the outfall showed slightly increased (~0.2 ppt) salinity near the substratum. Salinity was highest at Site 4, the prevailing current was from the north and could have directed the brine discharge towards this site. The 8th March profiles were relatively uniform although winds and swell in the morning were calm providing conditions, which could have favoured stratification. Waves were however slightly stronger Page

44 during the previous days. A series of profiles made around the outlet closer to shore failed to pick up increases in salinity. No discharge was however occurring at this time. Fig. 2 Salinity profiles at Binningup Site 3 to 7. Three repeated profiles recorded at each site. Site 5 orange profile represents a second series of measurements taken later on 8 March Dissolved oxygen Dissolved oxygen was 90% saturation and consistent throughout the water column, at Site 3 and Site 7, the reference sites 8 km and 5 km respectively from the Outfall (Fig. 13). Page

45 Fig. 3 Dissolved oxygen at Sites 3 Buffalo Rd through to Site 7 Myalup This pattern was also recorded at Site 6 Outfall North although there was a slight decrease in oxygen saturation on the 7th and 8th March. Sites 5 nearest to the outfall and site 4 south of the outfall also followed this pattern on 5th March. On 7th March at both sites there was a reduction in dissolved oxygen throughout the water column and a larger decreases towards the substratum, which was more evident at Site 5 where dissolved oxygen was reduced to approximately 60% saturation. On 8th March the % oxygen saturation was again higher in the upper water column (1-8 m), but still slightly lower near the bottom at m relative to the Page

46 % saturation. Observations of anoxic-like sediments were also noted by divers at site 5. No observation was made at Site 4. Profiles temperature and salinity adjacent to outfall March 2013 The locations of the CTD profiles area (Fig. 4) were immediately west (left side of the plot), and in line with the monitoring buoys that mark the Outfall restricted area. Temperature at all sites was higher, C at the surface, with an obvious thermocline (decreasing rapidly) between 2 and 4 m ( Fig. 5a). Temperature below the thermocline was C. Salinity ( Fig. 5b) was more variable with profiles at most locations indicating that the water column was well mixed, with a psu of 36. In general, the salinity profiles were similar to the other survey sites where CDT profiles were recorded on the 8 th March. However locations #s 8, 7 and 2, show low salinity water at 2 m, with a psu 34.7 at location #2. Below the thermocline (2-4m), salinity at locations #8 and #7 was higher (psu ) than the surface water, but slightly lower than the other seven locations, and the water column was well mixed. Salinity at location #2 however did not increase as markedly as salinity at locations # 7 and 8, and did not reach a similar well mixed profile until 6 m depth. Page

47 Fig. 4 Location of CTD profiles (1-12 m), adjacent to Site 5 Outfall 8 th March 2013 (1250 to 1315 hours). Temperature at Binningup outlet (a) Depth (m) Temperature ( C) Salinity at Binningup outlet (b) Depth (m) Salinity (psu) Fig. 5 (a) Temperature and (b) Salinity at Outfall Site 5. Locations were immediately west (left side of the plot), in line with the monitoring buoys that mark the Outfall restricted area (right hand side of plot).the coastline follows the right Y axis of the plot. Seagrass The ribbon weed, Posidonia angustifolia, was the predominant seagrass species, consistently present at all sites on sand substrate. Wire-weed, Amphibolis antarctica, was present at Sites 2 and 3 on sand covered reef pavement, occasionally with paddle weed, Halophila ovalis Page

48 growing under the leaf canopy of larger seagrasses or at the edge of patches (Sites 2, 3, 4). Posidonia denhartogii was also found at the northernmost Site 8. Seagrass biomass and leaf area were lower at sites closer to the outfall in both 2012 and 2013 but here were no significant differences in shoot density (Fig. 6, Fig. 7 and Fig. 8). In the 2012 survey, values for leaf biomass were lower at Sites 4-7 (34-48 g m -2 ), and higher at sites furthest from the outfall at Sites 1, 2 and 8 ( g m -2 ). Values for leaf area index (LAI) showed a similar pattern to biomass and ranged from approx m 2 leaf m -2 ground area. Shoot densities were variable reflecting the patchy growth form, and were highest at Site 2 (866 ± 77 shoots m -2 ) and lowest close to the outfall at Site 5 (450 ± 133 shoots m -2 ). Highly significant differences between years in biomass and leaf area index are attributed to differences in leaf length (Fig. 9), as there was no difference in shoot density between years (Table 5 and Table 6). In 2013, there was no seagrass present at Site 5 closest to the outfall and the sea bed was covered by a fine glutinous black (anoxic) substrate, possibly cm deep. Page

49 a) b) c) Fig. 6 Leaf biomass (a) Sites 1-8, 2012; (b) Sites 3-7, 2013; (c) Comparison of common sites sampled in both 2012 and Shared letters above bars indicate no significant difference (p>0.05) among sites. Page

50 a) b) c) Fig. 7. Leaf area index (a) Sites 1-8, 2012; (b) Sites 3-7, 2013; (c) Comparison of common sites sampled in both 2012 and Shared letters above bars indicate no significant difference (p>0.05) among sites. Page

51 a) b) c) Fig. 8. Shoot density (a) Sites 1-8, 2012; (b) Sites 3-7, 2013; (c) Comparison of shoot densities at common sites, sampled in both 2012 and Shared letters above bars indicate no significant difference (p>0.05) among sites. Page

52 a) b) c) Fig. 9. Mean length of the outer, mature leaves (a) Sites 1-8, 2012; (b) Sites 3-7, 2013; (c) Comparison of leaf lengths at common sites sampled in both 2013 and Shared letters above bars indicate no significant difference (p>0.05) among sites. Page

53 Table 5 Means and SE from six replicate quadrats for all seagrass measurements at each site in Site a Leaf biomass Mean (g/m2) Leaf biomas SE Shoot density Mean (shoots/m2) Shoot density SE Leaf area Index Mean (m2/m2) Leaf area Index SE Leaf length Mean (mm/leaf a) Leaf length SE Table 6. Means and SE from six replicate quadrats for all seagrass measurements at each site in Site 3 4 4a 6 6a 7 Leaf biomass Mean (g/m2) Leaf biomass SE Shoot density Mean (shoots/m2) Shoot density SE Leaf area Index Mean (m2/m2) Leaf area Index SE Leaf length Mean (mm/leaf a) Leaf length SE Page

54 Pavement limestone reef 2012 The area of reef in the study area was much lower than the area of seagrass and sand. Additionally the reef was very flat, of low profile, and regularly covered with sediment. Variability of the reef structure greatly influenced large sessile biota such as algae, sponges and corals, which in turn provided habitat for smaller less robust invertebrates. At Site 1 Peppermint Beach, the area of continuous rock was more extensive than the pavement at other sites. The surface was also higher than the adjacent non-vegetated sand and slightly elevated, cm, above the main pavement. Although algal cover was not quantified, algae were abundant, ranging in size from small species of low profile to large (> 10cm) upright foliose thalli. The algal community was also speciose including Rhodophyta, (Hypnea, Plocamium, Delisea, Pterocladia), Phaeophyta (Padina) and Chlorophyta (Codium spp. and Caulerpa). Depressions in the limestone surface contained sand, rock rubble and dead shells. The most notable of the large biota were upright (20-30 cm) sponges and the hard corals including the one metre high upright domes and plate-like Plesiastrea, Turbinaria, and Goniastrea and small colonies of favids. Large corals were only clearly visible at Site 1. At Site 2 Dalyellup, algal cover and the species present were similar to those at Site 1, however large thalli of the phaeophyte Scabaria were also present. Large sponges were prominent. At most other sites, the exception being Site 3 Buffalo Rd, sponges were generally small (<5cm). North of Bunbury the reef profile was increasingly lower and fragmented to the north and no continuous areas of rock were located for sampling at Site 7 and Site 8 the greatest distance north of the outfall. At Site 3 Buffalo Road, pavement in 2012 was slightly higher and continuous than at Sites 4, 5 and 6 nearest to the outfall. Similarly there was a gradient in algal cover and diversity along the coast with a greater variety of larger more abundant algae at Site 3 Buffalo Road and smaller, isolated thalli in lower areas with patchy reef. Invertebrate Diversity 2012 All species recorded were typical of species found along the south west coast, amongst seagrass, in sand or attached to seagrass leaves, and algae associated with limestone rock. Sessile Ascidians and Sponges (Porifera) dominated the invertebrate fauna. Sponges with 35 morpho-species, and 162 records were the most diverse (Table 7) followed by 17 Ascidians species with 364 records. Page

55 Sponges at the southern sites were predominantly large free-standing structures of various shapes -fan shaped, branched, mounds, spheres, vases or thin encrusting layers on the substrate. At the northern sites although some large sponges were present, small upright tubular forms in the family Sycettidae dominated. Ascidians were predominantly solitary species Polycarpa viridis, Pyura australis, Polycarpa clavata, Pyura sp. Colonial or compound ascidians of the family Holozoidae and Didemnidae were also common. Pyura australis and P. clavata were predominantly found on seagrass shoots. Echinoderms with 10 species from 5 Classes and 161 records were also prominent and included with five species (Asteroidea (sea stars), two species of Holothuria (sea cucumbers) and Echinoidea (urchins), and one species of Crinoidea (feather stars). Most records were single occurrences. The exceptions being the purple echinoid Heliocidaris erythrogramma, and the asteroid Meridiastra gunni, known previously as Pateriella brevispina. H. erythrogramma was particularly abundant on reef pavement at Site 1 (Peppermint Beach) with 77 individuals, Site 3 (Buffalo Rd) and Site 4 (Outfall North) with 34 and 13 individuals respectively. These were generally clustered in depressions in the pavement or in seagrass habitats adjacent to rubble and algae. The tests of dead H. erythrogramma, were also recorded at a number of additional sites and are likely to be present although perhaps not abundant, throughout the survey area. M. gunni was found amongst pavement (9 individuals) and seagrass (10) at Site 3. This relatively large (> 10 cm) mobile species was also recorded nearer to the outfall on pavement at Site 5 (2) and Site 6 (1) and in seagrass in seagrass at Site 4 (1). Page

56 Table 7 Total number of individuals (number of species) recorded at Sites 1-8, March and April 2012 Page

57 Molluscs, 10 species and 13 individuals, were represented by two bivalves, six prosobranchs and two opisthobranchs. These were all small species 2-3 cm in size, the largest being Campanile symbolicum (12 cm) at Site 1 Peppermint Beach. Dead shells recorded by divers as evidence of recent populations were generally representative of the species recorded in the quadrats but were not abundant or diverse at any site, the exception being depressions in the reef surface at Site 1 Peppermint Beach and Site 3 Buffalo Road. Three very large, recently dead carnivorous gastropods shells Syrinx aruanus (Site 6 Seagrass), Livonia nodiplicata and Melo miltonis (Site 3 Pavement) were also documented. Cnidaria, with 5 species included four stony corals Sclerectinea, one species of anenome (Actinaria) Heteractus malu, and a zoanthid Isaurus cliftoni, which with 36 individuals, generally solitary polyps, was the most abundant cnidarian. The large stony corals were only present on reef pavement at Site 1 Peppermint Beach. Small favid colonies were also recorded at Site 3 and Site 6 Outfall North, but were not abundant. Anemones H. malu were only recorded in sand areas within the pavement reef at Site 3 Buffalo Rd and in seagrass at Site 4. Other Actinaria and zoanthids e.g. Palythoa sp. were also observed on pavement at Site 3 but were not recorded systematically and not included in the analysis. Other groups included lace corals Bryozoa with five species (34 individuals). The most common bryozoan a Bugula species (27 records) was attached to seagrass and rock at all sites. One brachiopod species Magellania flavescens was found attached to pavement at Site 1 and 2 each with one record, and 12 individuals at Site 3. A few Crustacea were observed during sampling the most common being hermit crabs living in the dead shells however, they were not present within the quadrats. The low number of hermit crabs may reflect the generally low abundance of live and dead shells. Site Biodiversity 2012 The species, summarised in Phyla or Classes, and abundance varied greatly from site to site (Fig. ). In 2012 diversity was higher on pavement than in seagrass habitats. Forty taxa represented by 307 individuals were recorded on pavement at Site 3 Pavement whereas 13 species and 63 individuals were recorded in seagrass at that site. Similarly, pavement habitat at Site 1 Peppermint Beach, was represented by 20 species and 133 individuals and in seagrass by four species and 10 individuals. In pavement at Site 5 located nearest to the outfall 16 species (85 records), and Site 2 by 14 species and 42 individuals. In contrast, in Page

58 seagrass there were only five species and 17 individuals at Site 5, and at Site 2 nine species and 18 individuals. Ascidians were most abundant, 129 individuals in the pavement area at Site 3 Buffalo Rd. The most common species being the solitary ascidian Polycarpa viridis with 102 individuals, the colonial Holozoid Dispalia 14 individuals, and the encrusting black colonial Didemnidae (Fig. ). Porifera were most abundant in pavement areas of Site 2 and Site 3 and in seagrass at Site 4 immediately south of the outfall (Fig. ). Abundance of Porifera could be deceptive. At Site 4 a small Sycon sp. about 1 cm in height above the sediment were abundant, but some of these may be a single part of one colony, the base of which was buried in the sand. In contrast on pavement at Site 2 Dalyellup most of the sponges although less abundant were large 5-20 cm in width. Page

59 Fig. 20 Diversity of fauna (Phyla or Class) recorded at Sites Mean (± SE) of six replicate five meter transects in Seagrass and Pavement habitats. Page

60 2012 Multivariate analysis of invertebrate communities across sites Multivariate examination, Permutational Manova (PERMANOVA), of the fauna in 2012 at all sites, using the abundance of each species in seagrass and pavement at all sites indicated significant differences in seagrass and pavement habitats at all sites ( Table 8). CAP analyses were also significant (Trace statistics Q_m HQ_m: , P: and delta _1^2: , P: ; Fig. 10); showing a Cross Validation 53/84 = % correct among sites and habitats. The communities on pavement at Sites 1, 3 and 6 and in seagrass at Site 8 were distinct. Pavement at Sites 1 and Site 3 to the left of the CAP1 axis are widely separated from the seagrass communities to the right of the axis. The spread of pavement Sites 2, 4, 5 and 6 adjacent to the seagrass illustrates the similarity of the fauna across both habitats. Fig. 10 Species at all sites and habitats March-April Six replicate transects Sites 1-7 and two replicates at Site 8 Preston Lake South. Green pale symbols indicate seagrass and red dark indicate pavement. Page

61 Table 8 PERMANOVA table including all taxa sampled in 2012 across all habitats, first including all sites and secondly including just Sites all taxa Sites 3 to 7 - all taxa Source df MS Pseudo-F P(perm) df MS Pseudo-F P(perm) Ha Si HaxSi Res Total There were significant differences in communities associated with pavement habitats between sites (Trace statistics Q_m HQ_m: , P: and delta _1^2: 0.977, P:0.0001; Table, Fig. 11). Cross Validation analyses showed 35/36 = 97.22% correct. Pavement replicates within sites (Fig. 11) were consistent but there was some separation of Site 3 and Site 6 to the left of the CAP1 axis and Sites 1, 2, 4 and 5 to the right of the CAP1 axis. The fauna associated with Site 5 nearest to the outfall was unique and located in the middle of the axis. Page

62 Fig. 11 Species in pavement Sites 1-8 March-April Pavement areas were not located at Sites 7 and 8. Table 9. PERMANOVA table including all taxa sampled in 2012 at the two different habitats separately Seagrass - all taxa Pavement - all taxa Source df MS Pseudo-F P(perm) df MS Pseudo-F P(perm) Si Res Total Fig. 12 Species in seagrass habitats at Sites 1-8 in March-April 2012 Site 8 had only two replicates. Faunal communities associated with seagrass sites in 2012 were significantly different between sites (Trace statistics Q_m HQ_m: , P: and delta _1^2: , P:0.0001; Table, Fig. 12). Cross Validation technique showed 27/48 = 56.25% correct. Individual replicate transects within sites were however quite variable (Fig. 12). Reference Site 7 was the most distinct as was Site 8 the other northern (83%). Site 5, nearest to the outfall (33%) was located in the middle of the plot. Separation between the pavement fauna at Site 1 and Site 2 and to some extent pavement and seagrass at Site 2 from other pavement sites, provides a distinction between the site south of Bunbury and those north of Bunbury. Page

63 Fig. 13. Species in seagrass and pavement habitats at Sites 3-7 April CAP plot of sites adjacent to Binningup Outfall (Site 5). Green pale symbols indicate seagrass and red dark indicate pavement. Examination of the site north of Bunbury Sites 3 to Site 7, within a 13 km area of the Plant outfall indicated that there were significant differences between the seagrass and pavement habitats (Trace statistics Q_m HQ_m: , P: and delta _1^2: , P: ; Fig. 13) The CAP plot illustrated seagrass sites arrayed to the upper right of the plot and pavement site to the lower left. Cross Validation technique showed 33/54 = 61.11% correct. Cross validation indicated that pavement fauna at Site 3 (100%) and Site 6 (67%) were distinct and different from each other and from all other sites and habitats. The pavement community at Site 5 was grouped with the seagrass community at the same site and at Site 7 in the middle of the plot. The seagrass community at Site 5 was similar to communities in pavement at Site 4 pavement and seagrass at Site 6. Page

64 2013 Invertebrate Communities 2013 sampling Although the study in 2013 was focused on sites north of Bunbury closer to the Outfall, the species recorded were very similar to those present in However, species that defined the southern sites, the large stony corals and large sponges were not as common in the vicinity of the SSDC plant. Ascidians (15 morphospecies) and Porifera (nine morphospecies) (Table 10) dominated the fauna at all sites in Porifera were the most abundant with 1093 individuals compared to 187 ascidians. The next most abundant groups were the Bryozoa (46 individuals) followed by Zoanthidae (33) the Isaurus cliftoni species, Polychaeta (29), and Sclerectinidae (17), Echinoderms (16). Less than 10 individuals of Brachiopods (8), Molluscs (4), Actinaria (4) were recorded. A single decapod a hermit crab was found in a dead mollusc and a sea spider Pycnogonid. These were not however included in the data matrix as they were small, cryptic and found opportunistically. Porifera, sponges were abundant amongst pavement at Site 4 and Site 6, but lower in abundance in seagrass at these sites (Fig. 14). Porifera were particularly abundant on pavement at Site 6 with 500 individuals and at Site individuals. Porifera were least abundant in the seagrasses at the reference Sites 3 and 7, where Ascidians dominated. Ascidians were abundant at each site, but highest at Site 6 Pavement north of the outfall, and amongst seagrass at the reference Sites 3 and Site 7 at the greatest distance south and north of the Plant Outfall (Fig. 14). Ascidians, the largest being the solitary species Polycarpa viridus and Polycarpa clavata and an unidentified species, were represented by 28, 14 and 45 individuals respectively. The next most abundant species was the colonial Policlinidae (Aplidium). This species found near seagrass in Site 4 (9 zooids), Site 4 Pavement (1 zooid), Site 4a (9 zooids), Site 6 pavement (14 zooids) and Site 7 (three zooids). Some of the groups of small, 1-2 cm zooids, may, like the Porifera, have been part of a larger colonial individual. A number of Holozoid species including the large (6-9 cm) purple Sigillina (Holozoidae) were also relatively abundant, with 8 individuals, found at the three Site 6 transects in Seagrass/Pavement, Site 6a Seagrass, Site 6 pavement and Site 7 seagrass. Convoluted fan shaped colonies of another holozoid Sycozoa sp. with seven records was also present at Site 4 pavement. Page

65 Encrusting colonial ascidians included a black Didemnidae with 17 individuals, which were found as small colonies (5-10 cm) attached to firm substrate such as to rocks, algal holdfasts and the base of seagrass shoots. Echinoderms were represented by eight species and 16 individuals: two species of Holothuridea, Asteroidea, and Ophiuroidea, and a single species of Echinoidea and Crinoidea. Six species were recorded on pavement at Site 4 (Outfall South) but were generally represented by single individuals. These included two species of asteroid Meridiastra gunni and Goniodiscaster, each a single record. Two species of ophiuroids, three Clarkoma cf pulchra, one Ophiothrix sp. and three echinoids Heliocidaris erythrogramma were recorded. The two species of Holothuria Holothuria hartmeyeri and Stichopus cf mollis were also recorded at Site 4, but in seagrass together with the asteroid Goniodiscaster. Two Meridiastra gunni were also present amongst seagrass at Site 3 Buffalo Rd. No echinoderms were found at the outfall Site 5 near the diffuser array, which was covered by black ooze and decaying fragments of seagrass. The absence of echinoderms at Site 6, also in the outfall area is notable. Site 6 however had a greater abundance of Cnidarians, Actinaria (sea anemones) and zoanthids and small corals. Molluscs were only represented by 4 four species, each represented by one individual, Ostrea sp, in seagrass at Site 4a, Australium squamifera and Granata imbricata at Site 4 Pavement and Thalotia chlorostoma at Site 7 Seagrass. Table 10 Total of individuals recorded (Number of Taxa) at each site in March Page

66 Fig. 14 Diversity of fauna (Phyla or Class) recorded at Sites Note the scale on the y axis is highest at Sites 4 and 6 Pavement. Mean (± SE) of six replicate five meter transects in seagrass and pavement. Page

67 2013 multivariate analysis of invertebrate communities across sites In 2013 pavement reef habitat was only represented at Site 4 and 6. At Site 5 the absence of pavement and seagrass data was due to the low visibility and the accumulation of decomposing seagrass. Site 3 was covered with sand and no pavement was located. Evidence of large sand movement was also apparent nearer to the plant at Sites 4 and 6. Additional seagrass transects were surveyed at Site 4a. An additional pavement site, a mosaic of pavement, sand, seagrass, and rock rubble, is represented as Site 6a SP. Fig. 15 Species in seagrass and pavement habitats Sites 3-7 March Orange symbols in the centre of the plot represent Site 6a with transects in both seagrass and pavement habitats. There were significant differences between habitats and sites, and a significant interaction habitat and sites (Trace statistics Q_m HQ_m: , P: and delta _1^2: , P:0.0001; Table 11, Fig. 15). Cross Validation technique showed 30/46 = 65.21% correct. The plot illustrates a separation of taxa associated with seagrass and pavement, with pavement at Site 4 and Site 6 in the upper right and seagrass associated fauna to the left and bottom of the CAP1 and CAP2 axes. Seagrass habitats at the reference Sites 3 and 7 south and north and further from outfall are grouped together to the left of the CAP1 axis and the top of the CAP 2 axis. The outfall Sites 4 and 6 grouped together lie in the middle of the CAP1 axis. Site 6a SP where seagrass and pavement were present lies between the seagrass Page

68 and pavement groupings. The two seagrass habitat at Site 4 and Site 4a together with the seagrass at Site 6 form a tight group well separated from the seagrass grouping, at the reference sites Site 3 and Site 7 further from the outfall and away from the pavement sites. Table 11 PERMANOVA table including all taxa sampled in 2013 across all habitats all taxa Source df MS Pseudo-F P(perm) Ha Si HaxSi Res Total 45 Comparison faunal communities 2012 and 2013 The comparison of faunal surveys in 2012 and 2013 is confined to the area of particular interest, centered on a 13 km area Binningup Desalination Plant, from Site 3 Buffalo Rd in the south to Site 7 Myalup in the north. In 2013 sand movement affected pavement at Site 3, and no fauna was present. Sand inundation also limited selection of additional sites designed to increase statistical power in the analysis. Another significant change occurred at Site 5 located nearest to the outfall, where an area of seagrass and pavement with a fauna community similar to Site 4 and 6 in 2012, was either devoid of seagrass and fauna or hidden below a blanket of fine sediment and dead seagrass in Comparison of faunal communities in Sites 3-7 across all taxa, habitats, sites and years was significant at all levels (Trace statistics Q_m HQ_m: , P: and delta_1 2: , P:0.0001; Table 12, Fig. 16). The cross validation of factors showed 54/96 = 56.25% correct. The CAP plot shows a clear separation of seagrass sites to the right of the CAP1 axis, which are intermingled with pavement Sites 3, 4, 5, 6, surveyed in Also shown is a clear separation of 2013 pavement Sites 4, 6 and 6a (mosaic of seagrass and pavement) away from the 2012 sites to the left of the CAP 1 axis. Site 3 pavement, surveyed in 2012 is represented by a clear grouping (100%) of individual transects indicative of a distinct pavement community, the site however was buried by sand in 2013 and the fauna was not recorded. In contrast, the difference between seagrass transects in 2012 and 2013 is not distinct. At Site 3 seagrass in 2012 were only 16 % correct and are widely dispersed across the main axis of the CAP plot. Page

69 Table 12 PERMANOVA table including all taxa across Years, Habitats and Sites, as well as the interactions between all factors. All taxa Source df MS Pseudo-F P(perm) Ye Ha Si YexHa YexSi HaxSi YexHaxSi Res Total 95 Fig. 16 Fauna across Sites 3-7 in seagrass and pavement habitats in 2012 and Sampling in 2012 represented by solid symbols and 2013 by outline symbols. Red dark symbols indicate pavement and green pale represent seagrass. Site 3 pavement and Site 5 pavement and seagrass were not sampled. Page

70 Examination of all taxa in pavement habitat in each year and site was significant (Trace statistics Q_m HQ_m: , P: and delta _1^2: , P:0.0001; Table 13, Fig. 17). The Cross Validation technique showed 29/34 = 85.29% correct. The communities on pavement at Site 3, Site 5 and Site 6 in 2012 were the most distinct (100% correct) and each community was quite unique. Site 4 was also tightly grouped. In 2013 only two pavement habitats were surveyed. In 2013 pavement communities at Site 4 and Site 6 were more similar to each other, and located together and separately from the 2012 sites. Fig. 17 Pavement communities 2012 and No invertebrates were located on pavement at Site 3 or 5. Sampling in 2012 represented by solid symbols and 2013 by outline symbols. Comparison of fauna in seagrass sites were significantly different at all levels (Trace statistics Q_m HQ_m: , P: and delta _1^2: , P: ; Table, Fig. 18). The Cross Validation showed 30/62 = % correct. There was a distinct separation of individual seagrass sites, which were more dispersed than in pavement sites. The communities sampled in 2012 are located predominantly in the middle of CAP 1 axis. Site 5 is located in the plot at the top centre and close to Site 6. Site 6 communities in 2012 are placed Page

71 separately from those recorded in 2013, with the Site 6 (2012) associated with Site 5. In contrast the individual transects at Site and 2013 are placed closely together. Fig. 18 Seagrass communities 2012 and Sampling in 2012 represented by solid symbols and 2013 by outline symbols. Table 13 PERMANOVA tables analysing all taxa across Years and Sites. The two habitats, seagrass and pavement, are analysed separately. Seagrass - all taxa Pavement - all taxa Source df MS Pseudo-F P(perm) df MS Pseudo-F P(perm) Ye Si YexSi Res Total Ascidiacea and Porifera 2012 and 2013 Ascidiacea and Porifera (sponges) were the most diverse and abundant taxa recorded in the surveys 2012 and 2013 ( Page

72 Fig. ). The numbers of sponges across Sites 3-7 increased from 104 in 2012 to 1093 individuals in In contrast ascidians across Sites 3-7 decreased in abundance from 340 records in 2012 to 187 in Abundance of both groups was higher in pavement habitats. Sponges were more abundant in Sites 4 and 6 closer to the outfall. Although numbers of ascidians in transects within a site varied greatly, they were also more abundant in seagrass habitats Site 3 and Site 7 away from the outfall than at Site 4, 5 and 6. Fig. 30 Ascidians and Porifera in seagrass and pavement habitats Binningup 2012 and Note the difference on the scale of each graph. There were significant differences in the analysis of ascidians and sponges at all levels of the PERMANOVA ( Table 14 PERMANOVA tables analysing Ascidians and Sponges across Years, Habitat and Sites. Page

73 14) and CAP analyses (Trace statistics Q_m HQ_m: , P: and delta _1^2: , P :0.0001; Fig. 19) The Cross Validation technique showed 44/96 = % correct. Pavements at Site 3, 4, 5 and 6 sampled in 2012 were quite distinct and grouped to the right of the CAP 1 axis. Seagrass sites were all aligned to the right of the CAP 1 axis. The most consistent changes between 2012 and 2013 occurred in pavements at Site 4 and Site 6, which are grouped apart along the CAP1 axis from 2012 to A similar shift in 2013 was evident in some Site 4 seagrass transects. Site 3 and Site 5 pavements that in 2012 formed cohesive and separate groups could not be sampled in Fig. 19 Ascidians and Sponges in pavement and seagrass habitats. Sampling in 2012 represented by solid symbols and 2013 by outline symbols. Red dark Page

74 symbols indicate pavement and green pale represent seagrass. Orange outlines represent both habitats in The PERMANOVA ( Table 14 PERMANOVA tables analysing Ascidians and Sponges across Years, Habitat and Sites. 14) and CAP analyses of ascidians at all sites in 2012 and 2013 was significantly different at all levels (Trace statistics Q_m HQ_m: , P: and delta _1^2: 0.855, P:0.0001; Fig. ). The Cross Validation showed 26/96 = % correct. The most distinct groups were Site 3 pavement (100% correct), Site 6 pavement (83%) and Site 7 seagrass (67%). Site 4 pavement sampled in 2013 (67%) although scattered was separated from most other sites. Pavement communities in 2012 at Site 3 and Site 6 located to the left of the CAP1 axis were clearly separated from all other sites, however as pavement at Site 3 was not sampled in 2013 changes in the community are unknown. Page

75 Fig. 32 Ascidians in pavement and seagrass habitats. Sampling in 2012 represented by solid symbols and 2013 by outline symbols. Red dark symbols indicate pavement and green pale represent seagrass. Orange outlines represent both habitats in The PERMANOVA and CAP analyses of all sponges at all sites in seagrass and pavement were significant as well (Trace statistics Q_m HQ_m: , P: and delta _1^2: 0.855, P: ; Table 14 PERMANOVA tables analysing Ascidians and Sponges across Years, Habitat and Sites. Page

76 14 and Fig. 20). The Cross Validation showed 26/96 = % correct. Even though this low total percentage, sponges communities were very different at seagrass and pavement habitats and in some level from 2012 to In general, seagrass communities in both 2012 and 2013 formed more cohesive group than pavement sites. Fig. 20 Sponges 2012 and 2013 in pavement and seagrass habitats represented by solid symbols and 2013 by outline symbols. Red dark symbols indicate pavement and green pale represent seagrass. Page

77 Table 14 PERMANOVA tables analysing Ascidians and Sponges across Years, Habitat and Sites. 2.4 Discussion Little is known of the habitats and biota in the area north of Bunbury and the proposal to establish a desalination plant at Binningup was the impetus for this study. Concerns were expressed in 2010 that effluent from the plant would affect seagrass communities and the biota associated with them. Although seagrass and sand dominate areas reef pavement also occur, but the locations were unknown. In light of the absence of data about the dominant seagrass Posidonia angustifolia and no published surveys about the fauna this report documents for the first time invertebrates in the seagrass and reef habitats across eight sites spanning 70 km of coast over two years (2012 and 2013). The survey in 2012 also included mapping of seagrass and reef habitats for the first time. Delineation of habitats with towed video transects and divers in March and April 2012 at the commencement of the survey indicated that seagrass Posidonia angustifolia was the major habitat in the m depth range and present at all sites. Seagrass was generally surrounded by sand, which was also present as a thin veneer over adjacent patches of reef. The survey revealed a marked gradient of increasing sand depth and wave exposure from south to north over a 70 km length of coast. Reef habitat was\fragmented, with a marked decrease in height, complexity and extent to the north where seagrass and large areas of mobile sand were more extensive. Reef habitats at Peppermint Beach (Site1) and Dalyellup (Site 2) were more extensive, higher and of greater complexity than all other sites and supported diverse algal communities. Reef at Sites 3 Buffalo Beach, and the Outfall Sites 4, 5 and 6 further north, was progressively lower and flatter. Page

78 Observations of sand ripples at the northern sites, was also indicative of regular sand scouring and smothering. No pavement reef was located at Site 7 Myalup and Site 8 Lake Preston South. The original concept of sampling areas high relief was not achievable as rocky areas were very flat, barely above the height of the surrounding sand and more commonly referred to as reef pavement. As a result of the differences in reef height and the associated biotic communities at Site 1 and Site 2, these sites were not included in the 2013 survey, which focused on the Binningup area from Site 3 Buffalo Road to Site 7 Myalup. Seagrass Estimation of percentage seagrass cover was unreliable along this exposed coast due to the suspension of fine sediments and poor visibility, and this report is based on collected samples that provide details of leaf biomass, leaf length and shoot density, and calculations of leaf are at all sites. In 2012 there were clear differences in leaf biomass and leaf area between sites in the central area of the survey area, over a distance of approximately 10 km north and south of the SSDP pipeline, compared to more distant sites, km away. Leaf biomass and the leaf area per unit area of the seabed provided a useful indication of the abundance of seagrass. If there were approximately the same number of shoots per unit area, then differences between sites can be attributed to differences in sizes of plants, such as leaf length. The length of mature, outer leaves on the shoots determines the overall height of the leaf canopy. Leaf length is in turn determined by growth rates, which is strongly influenced by light availability. Turbidity in the water column decreases light reaching seagrass growing on the seabed, and much of the survey area appeared to be subject suspended sediment for most of the year. The question then arises are the lower biomass and leaf area values the result of impacts from development, such as pipeline construction and decades of dredge spoil dumping, or is this area more exposed to storm waves and sediment movement, so that the benthic habitats are more frequently disturbed resulting in smaller seagrass plant. There were also clear differences between years for the five sites sampled both in 2012 and In both cases, there were significant differences in shoot biomass and leaf area at any of the sites or between both years, which we attribute mainly to differences in leaf length. There is a seasonal cycle in species of Posidonia on the Western Australian coast, with a distinct autumn shedding of leaves that have grown long over summer. There are no data for P. angustifolia but it is likely to follow the seasonal pattern of the better-studied P. australis and Page

79 P. sinuosa (Cambridge & Hocking 1997). A likely explanation for the higher biomass and leaf areas recorded at several sites during the May 2013 sample is that the older, longer leaves had not yet been shed in 2013 resulting in significantly longer leaves (40-50 cm) compared to 20 cm or less in Invertebrate surveys 2012 Studies of invertebrate fauna associated with seagrass in Western Australia, has to date focused on small motile species epifauna macrofauna (> mm) associated with algal epiphytes that grow on the leaves and stems of seagrass or on similar fauna infauna within the sediment (Brearley and Wells, 1998 & 2000, Edgar 1992, Edgar & Robertson 1992, Gartner et al. 2010, Jernakoff et al. 1996, Lavery et al and Walker et al. 2000). A study in the southern section of Geographe Bay (Barnes et al. 2008) is one of the few focused on the larger motile and sessile fauna that can be counted and recorded in the field or collected only when voucher specimens are needed for identification. In the southern Geographe Bay study, the faunal groups documented (corals, porifera sponges, ascidians echinoderms and molluscs) were similar to those recorded at Binningup in this study. The two years of sampling provides the first assessment of species and communities in this area, and places the fauna in seagrass and reef pavement habitats in the context of other west coast habitats. Generally but not exclusively the same species were present in both habitats. Predominantly the species, within the framework of a number being undescribed were sessile filter feeders principally ascidians and sponges defined by morphology, shape, texture and colour, and a few mobile echinoderms that feed on sessile fauna or organic matter in the sediment. All were species previously observed along the southern west coast in both seagrass meadows and on reefs. Motile species included Echinoderms: Echinoids (urchins) Asteroids (sea stars), Crinoids (feather stars) and Holothurians (sea cucumbers), and Molluscs (gastropods). Generally these were common species but were recorded only once or twice. The exceptions being the purple echinoid Meridiastra gunni (previously Pateriella brevispina) at Site 1 Peppermint Beach and Site 3 Buffalo Road, and the purple-red sea star Heliocidaris erythrogramma at the reference Site 3, which were most common on reef pavement. The most abundant taxa were Ascidians and Porifera, which are sedentary filter feeders that attached to rocks, seagrass and dead shells. There were clear differences between the communities in seagrass and reef pavement habitats. The differences between sites were also significant, most noticeably due to the Site 1, Site 2 Page

80 Dalyellup and Site 3 in the south where species diversity and abundance was high and corals were present, and sites in the north where the abundance of fauna was lower. The fauna clearly reflected the higher and greater area of reef pavement in the south typified by Site 3 with the highest diversity, 40 species, and abundance 307 individuals recorded in pavement and 63 (13 species) in seagrass. In comparison Site 6, the most diverse site in the north had the highest number of records with 95 individuals (11 species) on reef pavement and 15 individuals (5 species) in seagrass. The abundance of invertebrates at Site 5 nearest to the Outfall in 2012 with 85 individuals (16 species) in pavement and 17 individuals (5 species) in seagrass was similar to Site 6. Invertebrate surveys 2013 In 2013 the survey focused on the sites in the vicinity of the SSD pipeline at Site 5, with Sites 4 Outfall 900 m to the south and Site 6 Outfall 300 m north, and reference Site 3 Buffalo Road 8 km south and Site 7 Myalup 5 km to the north. Extensive sand movement in late 2012 had blanketed the pavement at Site 3 and could be sampled. Similar scouring and smothering was also detected at other sites adjacent to the outfall where areas of reef pavement could not be located. At Site 5 no seagrass or pavement was located, and the area was covered with black ooze sand particles and fragments of seagrass (Figure 5 h-j). The disappearance of the black colouration when exposed to air an indication that the sediment sample was anoxic and seagrass was decaying. The species recorded were similar to those recorded in 2012 although with the focus on the northern sites, the overall diversity was lower. There were six additional species, and 20 species not recorded. Many taxa were in low abundance with only one or two records. Ascidians and Porifera (sponges), sedentary filter feeders dominated the fauna. Sponges were most abundant in pavement habitat and were in low abundance in seagrass at Site 3 and Site 7. In contrast ascidian numbers were higher at these sites. Notably, echinoderms were absent from Site 5 at the outfall and Site 6 north of the outfall. Comparison of fauna The multivariate analyses illustrated significant differences in the composition of fauna associated between years, sites and habitats (seagrass and reef pavement). Although many species were found in both habitats, the overall abundance in reef pavement areas probably reflects the stability pavement that also underlies sand adjacent to seagrass. The results were Page

81 dependent on all the taxa, but the differences between 2012 and 2013 also reflect the dominance of ascidians and sponges, which had different responses with sponge increasing and ascidians decreasing. Differences between habitats were largely driven not by the diversity of species but by the number of individuals. The higher abundances were in reef pavement and lower abundances in seagrass. Although a few species were attached to seagrass these were predominantly at the base of seagrass shoots. Temporal changes in the species presence and abundance are well documented and reflect the availability of larvae, successful recruitment, mortality and overall responses to the environment. The extensive movement of sand over reef pavement in late 2012 at Site 3 Buffalo Road illustrates a major decrease in that habitat, which in April 2012 had supported a diverse and abundant fauna predominantly sessile fauna. The few motile taxa were species that live on or within sand (molluscs), or species that can tolerate some sand cover or adhere to isolate areas of rock and seagrass (echinoderms). The dominance of sessile filter feeding taxa, ascidians and sponges, however is no doubt a reflection of water and sand movement on this exposed coast. The increase in sponges and decrease in ascidians could reflect different responses to the same conditions. The analyses also indicated significant differences in the communities at reference Sites 3 and 7, to those nearer to the Outfall (Sites 4, 5 and 6), and a similarity of the reference sites to each other. Overall sites nearer to the outfall were different to the reference sites further from the outfall. The highest increase in sponges occurred at sites nearest to the outfall, on reef pavement between 2012 and 2013 at Site 4 (0 to 387) and Site 6 (4 to 500), and in seagrass Site 4 (32 to 37) Site 6 (1 to128). A number of scenarios could have influenced the increase in sponges. Firstly the sponges may have been affected during the construction phase and or the early discharge of brine, prior to the 2012 survey, and responded by 2013 to improved conditions. The increase in sponges could also reflect particularly favourable conditions for sponge larval recruitment, establishment and growth over the 12 months between surveys. As high turbidity and suspension of organic matter can also favour some sessile filter feeders, higher sediment movement due to storms or Page

82 disturbance of the seabed during construction of the pipeline may have also favoured sponge growth between 2012 and Ascidians responses at different sites were inconsistent and confounded by different in pavement and seagrass habitats. The decrease in ascidians may reflect the high movement of sand prior to the 2013 survey. There was no pavement sampling at the reference Site 3 Buffalo Rd, due to sand smothering in 2013, but 129 individuals were recorded in Ascidian in the seagrass at Site 3 also decreased from 48 to 37, and sponges increased from three to seven individuals. Conversely at Site 7, the other seagrass reference site ascidians increased but sponges decreased. At the outfall sites, ascidians increased in pavement and seagrass at Site 4. In contrast at Site 6 ascidians decreased in both habitats. At Site 6 there were four sponges and 63 ascidians in There was however only a limited area of pavement and transects in 2013 covered a mosaic of seagrass and pavement, with only three ascidians but 24 sponges recorded. Despite the obvious effects of storm driven high sand movement, and associated increases in turbidity during the construction phase and excavations for the pipeline there are a number of other changes could have affected the invertebrate fauna. These include responses of the seagrass in the same period. In 2012 seagrass biomass and leaf area was lower at Site 4, 5 and 5a and 6 in 2012 than at the reference sites, that included Site 1 Peppermint Beach, 2 Dalyellup and Site 3 Buffalo Road in the south and north of the outfall at Sites 7 Myalup and Preston Lake South. In 2013 seagrass biomass was still lower at the outfall Sites (4, 4a and 6) than the reference Sites (3 and 7). Additionally in 2013 at Site 5 nearest to the outfall there was no seagrass and the area of pavement surveyed in both March and April 2012 (Figure 5 h-j). In combination, the lower seagrass biomass in 2012 and the disappearance of the seagrass at Site 5 provides some evidence of a continual decline in seagrass in the period between 2012 and 2013, and an indication that seagrass health in the outfall area was already compromised when the survey commenced. The absence of echinoderms in the outfall area is of particular interest. Echinodermata are the only stenohaline invertebrate phyla (Russell 2013), ie. they are osmoconformers that maintain an internal isosmotic environment and can only survive in a limited range of external osmolatities. In the case of most echinoderms the tolerance salinity range is very narrow (Ibid.). Sea urchin intolerance to salinity change has been noted at Monterey USA (Schlenk Page

83 pers. Com 2016), and in the Mediterranean (Raventos et al, 2006, Ferandez-Torquemada et al. 2013, de-la-ossa-carretero et al. 2016), where it was noted during routine desalination discharge monitoring that echinoderm populations disappeared during a period of high salinity release (in situ 38.2 psu), then recovered with salinity decline (in situ 36.9 psu) (Ferandez- Torquemada et al. 2013, de-la-ossa-carretero et al. 2016). The sensitivity of echinoderm populations to increase in in situ salinity levels is therefore useful as an early indicator to initiate remedial action; in the case cited above, the reduction of salinity supported the reintroduction of echinoderm populations and potentially prevented wider damage to key ecosystems such as P. oceanica meadows. On the exposed coastal area of Binningup, the water column is usually well mixed but stratification can occur during calm conditions, reducing mixing throughout the water column. During the 2013 invertebrate survey temperature, salinity, and dissolved oxygen at the three sites near the outfall and the reference sites 5 and 8 km north and south of the outfall, during three days of increasingly calm conditions that resulted in stratification. There was no brine discharge in this period. A short-term increase in temperature and salinity and lower oxygen near the seabed at Site 5 during the survey provides a snapshot of less than ideal conditions for biota during this short quiescent period. It is unlikely that the changes in the invertebrate community are due to a single factor but rather a number or series of conditions. The changes at the different sites could reflect a return to preconstruction conditions, which are unknown, or continued changes from 2012 to 2013 or represent natural fluctuations. There was however a complete loss of seagrass at the pipeline and reduced seagrass biomass a few hundred metres either side of the pipeline (Sites 4 and 6) that indicate a decline in the immediate vicinity of the outfall that could affect the seabed and nearby faunal communities. 2.5 References Anderson, M. J. Gorley, R. N. and Clarke K.R Permanova + for Primer: Guide to Software and Statistical Methods. Primer-E, Plymouth, UK. Barnes B. Westera M. Kendrick G. & Cambridge M Establishing benchmarks of seagrass communities and water quality in Geographe Bay, Western Australia. Annual Report to the South West Catchment Council. Bell J. J The use of volunteers for conducting sponge biodiversity assessments and monitoring using a morphological approach on Indo-Pacific coral reefs. Aquatic Conservation: Marine and Freshwater ecosystems 17: Page

84 Bell J. J. & Barnes D. K. A Sponge morphological diversity: a qualitative predictor of species diversity. Aquatic Conservation: Marine and Freshwater ecosystems 11: Bell J. J. Burton M, Bullimore B. Newman P. B. & Lock K Morphological monitoring of subtidal sponge assemblages. Marine Ecology Progress Series 311: Brearley A. & Wells F. E. (1998) Ecological significance of seagrasses: Task 8 Invertebrates. Shells and Dredging Environmental Management 79 pp. Brearley A. & Wells F.E. (2000) Invertebrate fauna in seagrasses on Success Bank, Western Australia. Biologia Marina Mediterranea. 7: Cambridge M. L. and Kendrick G.A. (2000). Habitat Mapping: Environmental Studies for Proposed Ocean Outlet, Bunbury Waste Water Treatment Plant. Department of Botany, The University of Western Australia in association with Alex Wyllie and Associates Pty Ltd, May pp. Clarke, K.R. Somerfield, P. J. and Chapman M. G On resemblance measures for ecological studies, including taxonomic dissimilarities and zero-adjusted Bray-Curtis coefficient for denuded assemblages. Journal Experimental Marine Biology and Ecology. 330: de-la-ossa-carretero JA, Del-Pilar-Ruso Y, Loya-Fernández A, Ferrero-Vicente LM, Marco-Méndez C, Martinez-Garcia E, Giménez-Casalduero F, Sánchez-Lizaso JL Bioindicators as metrics for environmental monitoring of desalination plant discharges. Marine Pollution Bulletin 103 (1-2): Edgar G. J Patterns of colonization of mobile epifauna in a Western Australian seagrass bed. Journal of Environmental Marine Biology and Ecology. 157: Edgar G. J. and Robertson A.I The influence of seagrass structure on the distribution and abundance of motile epifauna: pattern and process in a Western Australian Amphibolis bed. Journal of Experimental Marine Biology and Ecology. 88: Edgar G. J Australian Marine Life the plants and animals of temperate waters. Reed Books 544p Ferandez-Torquemada Y. Gonzalez-Correa, J.M. Sanchez-Lizaso J.L Echinoderms as indicators of brine discharge impacts Desalination Water treatment. Desalination and Water Treatment 51: Gartner A. Lavery P. S. McMahon K. Brearley & A. Barwick H Light reductions drive macroinvertebrate changes in Amphibolis griffithii seagrass habitat. Marine Ecology Progress Series. 401: Goudie L. Norman M. Finn J Sponges. A Museum of Victoria Field Guide to Marine Life. Museum of Victoria. ISBN: (e-pub) , (e-pdf) Gowlett-Holmes K A field guide to the marine invertebrates of South Australia, Notomares pages. Jernakoff, P., Brearley, A. & Nielsen, J. (1996) Factors affecting grazer-epiphyte interactions in temperate seagrass meadows. Oceanography and Marine Biology: an annual review. 34: Keesing J. K. and Heine J. H. (eds.) 2006 Strategic Research Fund for the Marine Environment (SRFME) Final report. CSIRO Marine and Atmospheric Research. Page

85 Kendrick G. Coupland G. McDonald J. Chatfield B Characterising the marine benthic habitats of the proposed southern seawater desalination Plant (SSDP) site: Interpretation from underwater towed video and map interpolation. Report to KBR Australia by the Marine Research Group University of Western Australia. Kott K The Australian Ascidiacea Part 1 Phlebobrnachia and Stolidobranchia. Memoirs of the Queensland Museum 23: Kott K The Australian Ascidiacea Part 2 Aplousobranchia. Memoirs of the Queensland Museum 29: Kott K The Australian Ascidiacea Part 3 Aplousobranchia (2). Memoirs of the Queensland Museum 32: Part Kott K The Australian Ascidiacea Part 3 Supplement 2. Memoirs of the Queensland Museum 32: Part Kott P Tunicate in Shepherd S.A. and Davies M. (1997) (Eds.) Marine Invertebrates of Southern Australia Part III. South Australian Research and Development Institute (Aquatic Sciences in conjunction with the Flora and Fauna of South Australia Handbooks Committee Lavery, P., Vanderklift, M., Hyndes, G. & Brearley, A. (2000a) Ecological significance of seagrasses Flora and fauna diversity Phase 3 Report Shellsand Dredging Environmental Management Programme 68 pp. Poore, Gary. C. B Marine Decapod Crustacea of Southern Western Australia. A Guide to Identification. Museum of Victoria ISBN: Raventos, N. Macpherson E. Garcia-Rubies A Effect of brine discharge from a desalination plant on macrobenthic communities in the NW Mediterranean. Marine Environmental Research (62) Russell, M.P 2013 Echinoderm responses to variation in salinity. Advances in Marine Biology. 66, Smit, A.J., Brearley, A., Hyndes, G.A., Lavery, P.S. & Walker, D.I., (2005) Carbon and nitrogen stable isotope analysis of an Amphibolis griffithii seagrass bed. Estuarine Coastal and Shelf Science. 65: Smit, A.J., Brearley, A., Hyndes, G.A., Lavery, P.S. & Walker, D.I., (2006) δ 15 N and δ 13 C analysis of a Posidonia sinuosa seagrass bed. Aquatic Botany. 84: Walker, D.I., Kendrick, G.A. Brearley, A, Lavery, P. Connell, S. Lantzke, R. Annandale, D. Wells, F.E. & Hillman, K. (2000) Ecological significance of seagrasses Phase 4 Final Synthesis Report 187 pp. Page

86 Page

87 3. Biomarker development Seagrass Morphological and physiological effect of elevated seawater salinities and desalination-derived brine on key seagrass species. Marion Cambridge 1, John Stratton 1, Andrea Zavala-Perez 1, Renae Hovey 1, Greg Cawthray 1, Julie Mondon 2 1.Oceans Institute, School of Plant Biology, University of Western Australia 2.Center for Integrative Ecology, School of Life and Environmental Sciences, Deakin University Key questions 1. To what extent does elevated salinity and desalination-derived brine exposure result in a deleterious stress response? 2. Which response, or suite of responses, are most suitable as biomarkers of exposure and effect from elevated salinity? 3.1 Overview A series of physiological and morphological responses in seagrass are evaluated in relation to whether they constitute valid biomarkers of effect from elevated salinity and brine exposure. Elevated salinity is a well-known stressor for most plant species. Desalination brine consists of concentrated sea salts plus other additives depending on the processes. The primary objective of this series of investigation is to identify and quantitatively evaluate the responses to increasing salinity at varying levels of seagrass maturity under experimental conditions. Two species were investigated as model organisms to identify representative salinity response in common seagrass species, enabling identification of Dose-response alteration Suitability for replication and in situ evaluation, and Model test species and testing procedures for future studies A comprehensive suite of response endpoints tested include changes related to survival, growth, photosynthesis, metabolism and osmoregulation, osmotic pressure, and carbohydrate and amino acid concentration. NB. Salinities are given as practical salinity units (psu), approximating the older usage of ppt: parts per thousand. Seawater salinity is approx. 35 psu, or 3.5% salt by weight. Page

88 Experimental suite summary Pilot study This testing procedure investigated the effect of one concentration of elevated salinity, 42 psu (practical salinity units) corresponding to approx. 25 % desalination brine diluted with ambient seawater. Two test species were exposed to elevated salinity for 4 weeks; Posidonia australis a large slow-growing seagrass, expected to have a limited capacity to adjust to increased salinity, and Halophila ovalis, a small, fast-growing seagrass expected to cope with higher salinity. The larger Posidonia australis exhibited greater consistency in response and was selected as the model test species for further experimentation. Elevated seawater salinity exposure This testing procedure investigated the exposure effect of 3 salinity concentrations (37, 46, 54 psu) corresponding to ambient seawater, approx. 50% desalination brine and undiluted 100% SSDP brine, on Posidonia australis shoots over 4 weeks. Significant response to elevated salinity was evident within 2 weeks at 54 psu. Survival was not significantly affected at intermediate salinity corresponding to approx. 50% dilution (46 psu), however metabolic and root growth inhibition was detected. Mature seagrass brine exposure This testing procedure investigated desalination brine (54 psu) on seagrass shoots to compare effects of brine exposure verses elevated seawater salinity. This experiment examined survival, growth, photosynthesis and osmoregulation (osmolality) to assess whether other components that may have been concentrated or added during the reverse osmosis process could produce a different response to elevated seawater salinity. The exposure duration consisted of a 2 week exposure at 25%, 50% and 100% brine (42 psu, 46 psu and 54 psu respectively). Leaf growth, turgor pressure, ionic concentrations changed with increasing brine concentration. Seagrass seedling brine exposure This testing procedure investigated the effect of brine on germinating seeds over a longer term exposure of 50 days. The rationale for investigating responses at the earliest stage of the seagrass life cycle experiment, is based on mature seagrass having exhibited a degree of short term tolerance to elevated seawater salinity and brine exposure, but significant decrease in leaf and root development in seedlings occurred at 50% and 100% brine exposure within 2 weeks. Page

89 3.2 Pilot study Plant collection and preparation Posidonia australis and Halophila ovalis plants were collected in the field from a shallow meadow m deep at Woodman Point near Fremantle. Posidonia australis transplants consisted of two shoots (with one apical shoot, and a cm length of rhizome) and Halophila ovalis was selected based on four mature shoots with one apical shoot, connected by underground stem internodes (rhizomes). Excavated plants were placed in mesh bags underwater, cleaned of sediments, then transported in seawater in an insulated carrier and transplanted into the culture tank system within 3 hours of collection. Tanks were placed in indoor laboratories to allow culture with constant background conditions (temperature, light, day length) with adjustments to salinity of natural seawater using aquarium salt. Figure 1: Diagrammatic representation of Posidonia australis and Halophila ovalis component parts. Experimental design Seagrass plants in culture were used for experiments on raised seawater salinity in large aquaria at aquaculture facilities (Department of Fisheries WA) to study the natural osmoregulatory capacities, as well as the capacity of basic metabolic functions to withstand raised salinity. Posidonia australis shoots were collected in the field from a shallow meadow m deep at Woodman Point near Fremantle. To ensure each rhizome division was cultured under Page

90 standardized conditions, each nursery tray was filled with a standardized sediment type; calcareous sediment with a heterogeneous grain size distribution and organic matter (dried seagrass leaves) added at 1.5% sediment dry weight. Calcareous sediments were sourced from marine dredging of seagrass habitat in Cockburn Sound, south of Perth, Western Australia. Organic content was determined by combusting 20 g of oven-dried sediment, for 4 hours at 450 o C. Combusted sediments were then weighed and percent organic content was calculated. Six P. australis and six H. ovalis transplants were planted into each tray and two trays (one of each species) were assigned to each tank (eight tanks). Each rectangular (100 cm L x 80 cm W x 60 cm D) 200 L tank received filtered (11 µm) seawater (19-21 o C) sourced offshore from the facility at a rate of 2 L min -1. The tank system was a semi-recirculating, where water was changed (30% total volume) every two days. A single 250 Watt metal halide light with reflective hood was mounted above each tank (Fig. 2). Plants were acclimated in tank culture for two weeks at ambient salinity (37 PSU, Practical Salinity Units) and a light intensity of 120 µmol m -2 s -1 (photoperiod of 10 hours light and14 hours dark) which represented in situ light conditions at the location and time of plant collection. Salinity and temperature were logged daily using a TPS logger and Light intensity (Photosynthetically Active Radiation, or PAR, units µmol m -2 s -1 ) was measured every two days using a LI1400 2π quantum logger with PAR sensor (190 series) to ensure salinity, temperature and light levels remained the same over the acclimation period. After two weeks acclimation, the number of shoots per tray was counted for each species for each tank. After an initial acclimation period the salinity level was elevated to 42 ppt in four replicate tanks. Each tank held two nursery trays containing six transplanted rhizome sections of each of the test species. Measurements were then made on seagrass plants to determine their physiological and growth responses to salinity. Survival over the experimental period was observation-based and identified by the presence of above-ground healthy shoots for each species under each salinity treatment. Death was identified by the absence of above-ground shoots for each species under each salinity treatment. Page

91 Fig. 2 Experimental mesocosms consisting of three 200 L tanks per treatment, overhead lighting and semi-recirculating flowing seawater for culturing seagrass plants under controlled conditions. Shoot growth and survival Total shoot counts per tray were measured before installation of salinity treatments (time 0) then at the end of the experimental period 30 days after installation of salinity treatments. Leaf lengths were measured before installation of salinity treatments (time 0) then and again after 30 days after treatment application. Posidonia australis leaves were measured from the leaf sheath to the tip of the leaf (see Fig. 1). Halophila ovalis leaves were measured from the base of the petiole to the tip of the oval-shaped leaf (see Fig. 1). Increase in rhizome length was assessed by measuring the distance between the final apical shoot (see Fig. 1) and the initial apical shoot present before installation of salinity treatments. That is, rhizome length was measured at 0 and 30 days. Measures of root length were conducted on harvested rhizome sections. Survival over the experimental period was observation-based and identified by the presence of above-ground healthy shoots for each species under each salinity treatment. Death was identified by the absence of above-ground shoots for each species under each salinity treatment. Page

92 Biomass Biomass was measured on harvested rhizome sections. After 30 days under experimental treatments, all rhizome sections of each species were harvested from each tank. Harvested plants were separated into their component parts (leaves, leaf sheaths, roots and rhizomes) dried in an oven at 60 C for 72 hours and weighed. Photosynthesis using chlorophyll fluorescence Chlorophyll fluorescence measures were conducted at the beginning of the experiment, and then at weekly intervals, using a pulse-amplitude modulated fluorometer (Diving-PAM fluorometer - Walz GmbH, Effeltrich, Germany). The potential quantum yield of photochemistry (Fv/Fm) was evaluated by subjecting dark-adapted leaves (dark adapted for 15 minutes) to a saturating pulse of light. The potential quantum yield of photochemistry (Fv/Fm) is typically used as a measure of stress in plants. Fv/Fm is calculated from the equation (Fm Fo)/Fm and thus is derived from Fo, initial dark-adapted fluorescence and Fm, the maximum dark-adapted fluorescence. Chlorophyll fluorescence parameters (Fo, Fm) were measured on 4 replicate leaves (mature, fully expanded leaves) for each replicate salinity treatment for each species weekly after installation of salinity treatments. The fiber-optic probe was held in place in a dark-leaf clip (Walz, Diving leaf clip) half way along the long axis of the leaf for each species. Fo (dark-adapted state) was determined by a weak pulsed red light (<1 µmol quanta m - 2 s -1 ), and Fm (dark-adapted state) was determined by a saturating pulse of white light (0.8 s at 8000 µmol quanta m -2 s -1 ). Leaf osmotic potential, water potential and turgor pressure Osmotic (Ψπ) potential were measured using a Fiske freezing point Osmometer, analyzing four leaves harvested weekly from each species and each replicate tank. Units are expressed as a negative pressure, mega pascals (MPa). After acclimating plants for 2 weeks in ambient seawater salinity, the salinity for the treatment plants was raised to 42 psu for 2 weeks. The water potential of ambient seawater at 37 PSU is -2.6 MPa and -3.1 MPa at 42 psu. Pilot Study Results All plants survived in control and raised salinity conditions during the experiment. Leaf growth in Posidonia australis showed a small, non-significant reduction in leaf growth at raised salinity compared to controls. Root and rhizome length did not differ significantly over Page

93 the test period (Fig. 3a). This species is extremely slow-growing so there was no change in shoot number. Fig. 3a. Increase in leaf length per Posidonia australis plant (n.s., n = 24) over 2 week test exposure at control (37 psu) and raised salinity (42 psu). Halophila ovalis increased the number of shoots during the test period, reflecting its fast growth rates but there was no significant difference at raised salinity compared to controls. However, the size of leaf was significantly smaller at higher salinity (Fig. 3b). Fig. 3b. Differences in leaf growth of Halophila ovalis at control (37 PSU) and raised salinity (42 PSU) shown as size of mature leaves formed during 2 week test exposure (p < 0.001, n=24). Page

94 Photo-physiology There were no significant differences in quantum yield for either species of seagrass with exposure to higher salinity (Fig. 4a, b) Fv/Fm Time 1: 26/10/2012 Time 2: 31/10/2012 Time 3: 5/11/2012 Time 4: 9/11/ Treatment Fig. 4a. Quantum yield (Fv/Fm) of Posidonia australis leaves measured at 4 day intervals during exposure to control (37 psu) and higher (42 psu) salinity. (n= 24) Photosynthesis Halophila ovalis Fv/Fm Treatment control 36 PSU, high salinity 42 PSU Time 1: 26/10/2012 Time 2: 31/10/2012 Time 3: 5/11/2012 Time 4: 9/11/2012 Fig. 4b. Quantum yield (F v/f m) of Halophila ovalis leaves measured at 4 day intervals during exposure to control and higher salinity. (n= 24) Page

95 Leaf osmotic potential, water potential and turgor pressure For Posidonia australis, mean osmotic potential in the leaves decreased from -3.29±-0.23 MPa in controls (ambient 37 PSU) to -3.57±-0.34 (treatment 42 PSU). Over the next 2 weeks, measurements were made at 4 day intervals. Osmotic potentials of control plants remained less negative than the treatments, with the final reading of -3.21±-0.27 (controls) and -3.37±-0.16 (treatment) but this difference was not significant (Fig. 5a). In each case the osmotic potential of the leaves remained more negative than the water potential of the external seawater medium. Fig. 5a Osmotic potential of Posidonia australis leaf tissue (mean±s.e.), after plants were cultured at ambient seawater salinity (37 psu), then exposed to raised salinity (42 psu) for 2 weeks. Osmotic potential responses in seagrass leaf tissue after 4 weeks exposure to raised salinity, n= 24. No significant change in Posidonia australis, Halophila ovalis significant, p< For Halophila ovalis, results were more variable, beginning with a small difference between control plants (-3.49±-0.3) and treatment at 42 PSU (-3.59±-0.3). By the end of the experiment, osmotic potential of plants at the higher salinity had decreased to -3.88±-0.16 MPa, whereas the controls were far less negative, -2.89±-0.21, but still more negative than the external seawater medium (Fig.5b). Page

96 Fig. 5b. Osmotic potential of Halophila ovalis leaf tissue (mean±s.e.), showing a weak significant difference (p<0.05) after plants were cultured at ambient seawater salinity (37 psu), then exposed to raised salinity (42 psu) for 2 weeks, n = 24. Summary of findings Pilot raised salinity exposure to Posidonia australis and Halophila ovalis Trend towards reduced leaf growth in P. australis at higher salinity Reduction in leaf growth in H. ovalis at higher salinity Decrease in osmotic potential in P. australis at higher salinity Decrease in osmotic potential in H. ovalis at higher salinity H. ovalis exhibits less consistency in response relative to P. australis Page

97 3.3 Elevated seawater salinity tolerance trials This experiment investigates the effect of exposure on seagrass shoots to 3 levels of salinity (37, 46, 54 psu), corresponding to ambient seawater, approx. 50% desalination brine and undiluted 100% brine, over 4 weeks and 6 weeks duration. Methods Tank system and experimental design The mesocosm system consisted of nine 200 L tanks with 3 tanks per treatment (Figure 6). Each treatment consisted of a closed seawater circuit feed from a 1000 L reservoir tank. Water was circulated using a 2500 l h-1 pump and drained back into the reservoir tank continuously. Artificial light ensured a consistent light quality simulating natural summer conditions, provided by one overhead Sylvania BT28 Metal Halide 250 W lamp (Osram Sylvania Inc, Danvers, MA, USA) per tank, set to 100 ± 20 μmol m-2 s-1, measured with a LI-1400 Data Logger (LI-COR Inc., Lincoln, NE, USA) spherical quantum sensor in a 12 hour light:dark cycle. Water temperatures in the tanks ranged between 21 and 26o C (mean 23.9 ± 0.8o C). The plants were held in the tanks for 5 days before commencing the experiment in natural seawater with a salinity of 37 psu-practical salinity units (range 36.4 to psu), which is the same as salinity in the area where plants were collected. Salinity treatments were imposed without a period of gradually increasing salinity, in order to simulate sudden increments in salinity associated with a brine discharge. Natural seawater was used for controls. Salinity was raised in two treatments: to 46 and 54 psu by adding artificial marine salt (Aquavitro, Seachem, Wilmington, USA) to the reservoir tanks until the desired salinity was reached. These salinity levels were maintained for six weeks until the end of the experiment. Salinity and temperature were monitored daily using a multiparameter field logger TPS 90-FLT (TPS P/ L, Springwood, Brisbane, Australia). All other experimental details were the same as for the pilot study. Page

98 Experimental design SW 37psu 1 46 (48)psu 1 54(57) psu 1 SW 37psu 2 46(48)psu 2 54(57) psu 2 SW 37psu 3 46(48)psu 3 54(57) psu 3 11 plants Each plant with 2 shoots Fig 6. Conceptual diagram of experimental design for salinity testing of Posidonia australis plants. Three tanks per treatment were used for experiments, with 6 weeks exposure to raised salinity 46 and 54 psu concentrations and seawater controls, 37 psu. Salinity increases over 3 days shown in brackets. Survival and Growth Survival over the experimental period was observation-based and identified by the presence of above-ground healthy shoots for each species under each salinity treatment. Death was identified by the absence of living above-ground shoots for each species under each salinity treatment. Leaf and root tissue were measured every two weeks on three randomly selected harvested plants. Leaf growth rates were measured following the method by Short and Duarte (2001). Seagrass leaves were marked 5 to 7 days prior to harvesting. After harvesting new tissue was measured to estimate shoot leaf growth rate (cm 2 shoot -1 day -1 ). Completely new leaves (leaves with no marks) were counted and measured entirely as new tissue. Additionally, root lengths were measured on harvested plants and growth rates (cm shoot -1 day -1 ) were estimated. Chlorophyll fluorescence and photosynthesis Chlorophyll fluorescence measures were conducted using a pulse-amplitude modulated fluorometer (Diving-PAM fluorometer - Walz GmbH, Effeltrich, Germany). The potential Page

99 quantum yield of photochemistry (Fv/Fm) was evaluated by subjecting dark-adapted leaves (dark adapted for 15 minutes) to a saturating pulse of light. After dark adaptation, all reaction centers of photosystem II are open and heat dissipation is minimal, therefore the maximal quantum yield of PS II could be observed. Measurements were carried out in the morning before lights were on to ensure complete dark adaptation; four randomly selected leaf replicates were measured per tank (in total 12 replicates per treatment); in order to minimize the associated leaf age variation. The fiber-optic probe was held in place in a dark-leaf clip (Walz, Diving leaf clip). Clips were placed on each leaf at the same distance from the base (approx. 3 cm). Chlorophyll fluorescence measurements consisted of a saturation pulse that induced maximal fluorescence yield (Fm) and maximal variable fluorescence (Fv = Fm - Fo). Fv/Fm indicates the potential quantum yield of PS II. The potential quantum yield of photochemistry (Fv/Fm) is typically used as a measure of stress in plants. Fv/Fm is calculated from the equation (Fm Fo)/Fm and thus is derived from Fo, initial dark-adapted fluorescence and Fm, the maximum dark-adapted fluorescence Fo (dark-adapted state) was determined by a weak pulsed red light (<1 µmol quanta m -2 s -1 ), and Fm (dark-adapted state) was determined by a saturating pulse of white light (0.8 s at 8000 µmol quanta m -2 s -1 ). Changes in Fo and Fm were used as stress indication (especially an increase in Fo). Measurements were performed at the beginning and every two weeks until the end of the experiment. Rapid Light curves (RLC) were performed in four randomly selected leaf replicates per tank (in total 12 replicates per treatment) at the beginning and every two weeks until the end of the experiment. Leaves were clipped at the same distance from the base (approx. 3 cm) as in the chlorophyll fluorescence measurements. RLC provide an insight on the induction and saturation characteristics of photosynthesis. Four parameters were estimated: Photosynthetic quantum efficiency (α), Maximum Electron Transport Rate (ETRmax), Minimum saturating irradiance (Ralph and Gademann 2005) or onset of light saturation (Schwartz et al. 2000) (Ik or Ek) and Saturation irradiance or optimal light intensity (Iopt. or Em). Water relations Osmotic (Ψπ) and water potentials (Ψw) were measured from three randomly selected harvested plants using a Dewpoint Potential Psychrometer WP4. From each replicate plant a leaf segment (approx. 4 cm long) was cut and placed into the Psychrometer chamber to measure osmolality. Page

100 Leaf segments were cut at a standardised distance from the base (approx. 3 cm) as it has been demonstrated that seagrass leaves maintain an osmotic pressure gradient along the leaves (Tyerman et al. 1984, Sandoval et al. 2012). In each replicate, osmolality was measured in fresh tissue (water potential) and frozen tissue (osmotic potential) according to Tyerman (1982). Fresh tissue was kept submerged during handling to minimize evaporation; then quickly blotted dry before placing into the chamber. After water potentials were measured, sample tissues were put into liquid nitrogen to freeze immediately. Out of the liquid nitrogen, sample tissues were placed into the chamber cup, then covered until the chamber temperature was equilibrated and osmotic potentials were measured. Leaf turgor pressure was calculated as the absolute difference between water and osmotic potentials (Kramer and Boyer 1995). Ion concentrations Inorganic cations (Na +, K + and Ca ++ ) and Cl - anions were determined in leaf tissue from the remaining nine plants in each tank at the end of the experiment. Leaves were cleaned to remove epiphytes and salts on the leaf surface, freeze-dried and ground to a powder in a ball mill grinder. Subsamples of the ground material were dissolved in 5 ml of 0.5M HCl for Na + / K + analysis and 20 ml of boiling DI water for Cl - analysis, and shaken for two days. Leaf tissue cation concentrations (Na +, K + and Ca ++ ) were determined using an ion chromatograph by flame photometry (Corning, Model 410, Halstead, Essex, UK). Anion concentrations (Cl - ) were determined with a boiling water extraction of ground tissue, followed by chlorometric impulse titration with a chlorinity meter (Model , Buchler Instruments, New Jersey, USA). Sugars and amino acids - Extraction The non-structural carbohydrates (NSC) of glucose, fructose and sucrose and free amino acids (FAA) were determined in freeze dried and finely ground leaf and rhizome tissue obtained from plants collected in the experimental units and stored at -80 C. The method of extraction for NSC and FAA was adapted from Marín-Guirao et al. 2013, Sandoval-Gil et al. (2013) and Warren (2000): 20 mg of ground dried sample was extracted with 1.0 ml of 95% (v/v) ethanol at 80 C for 15 min, and then centrifuged. After removal of the supernatant, the pellet was resuspended and the extraction repeated. The supernatants were combined to provide the extracted sample for NSC and FAA analysis as detailed below. Page

101 For FAA analysis only, aliquots of 700 µl for leaf and 500 µl for rhizome were taken to dryness under vacuum with a SpeedVac (Savant, Farmingdale, NY, USA) and subsequently re-dissolved in 150 µl Milli-Q water containing 56.6 µm of the internal standard, α-amino-nbutyric acid. Soluble sugars - Analysis Analysis of NSC was adapted from Slimestad (2006) and undertaken using a high performance liquid chromatograph (HPLC) consisting of 600E pump, 717plus autoinjector, (Waters, Milford, MA, USA) and an Evaporative Light Scattering Detector (Alltech 500 ELSD, Grace Materials Technologies, Deerfield, IL, USA). Separation was achieved at 30 ± 0.5 ºC on a Prevail ES Carbohydrate column (250 x 4.6 mm i.d. with 5 µm packing; Grace Materials Technologies) using an isocratic mobile phase consisting of 25% Milli-Q water and 75% acetonitrile at 1 ml min -1. All solvents were vacuum filtered to 0.22 µm prior to use and were continually degassed with helium sparging Samples in the autoinjector were held at 10ºC and the ELSD drift tube was held at 80ºC with high purity nitrogen flow rate of 2.5 litres per minute for nebulisation. Calibration curves for each sugar were generated from ELSD peak area (log10) versus the mass of standard sugar injected (log 10), and a standard analysed every 10 samples to check for any instrument/detector drift. Data acquisition and processing was with Empower 2 (Waters Corp., Milford, Massachusetts, USA) software. Retention times of sugar standards were used to identify sugars in the sample extracts. Typical sample injections were 10 µl and runtime was 15 min per sample. Free amino acids - Analysis Analysis of FAA was undertaken using the Waters AccQ.Tag chemical package for HPLC, with a Waters system consisting of a 600E dual head pump, 717 plus autosampler, a 470 scanning fluorescence detector and a 996 photodiode array (PDA) detector. Separation was performed on a Waters Nova-Pak C18 column (150 mm x 3.9 mm I.D.) with 4 µm particle size, held at 35 C. All data were acquired and processed with Empower chromatography software (Waters) with fluorescence detector settings of 250 nm Excitation; 395 nm Emission, 0.5 filter, 100 gain and the PDA set to 250 nm. Positive identification of amino acids was accomplished by comparing standard retention time for fluorescence and PDA as well as PDA peak spectral analyses with the samples. Page

102 For sample derivitisation, a 10 µl aliquot of the re-dissolved extract 70 µl of the kit-supplied borate buffer was added and samples vortex mixed. Then 20 µl of the AccQ.Fluor reagent was added and samples immediately vortex mixed. Samples were then incubated at 55 C for 10 min prior to HPLC analysis. Standards and blanks were carried through the same procedure. A gradient mobile phase system was employed with reservoir A consisting of the Waters AccQ.Tag aqueous buffer, prepared from the concentrate following manufacturers instructions supplied with the kit, reservoir B was acetonitrile and reservoir C was Milli-Q water. All solvents were vacuum filtered to 0.22 µm prior to use and were continually degassed with helium sparging. The gradient profile was as per the instructions with the kit. The mixed amino acid standard contained aspartic acid, glutamic acid, glycine, arginine, threonine, alanine, proline, γ-aminobutyric acid, cysteine, tyrosine, valine, methionine, lysine, iso-leucine, leucine, phenylalanine, tryptophan, as well as amino acid pairs, serine+asparagine and histidine+glutamine, which could not be separated with the HPLC method used. Data analyses All the studied parameters tested the null hypothesis that there were not significant differences among salinity treatments or sampling times. Analyses consisted of nested analysis of variance (ANOVA) design with Treatment and Time as fixed factors, while Tank was treated as random factor nested within Treatment. Data were transformed (log(x+1)) when necessary and ANOVA assumptions were tested. If data presented significant deviations from normality and homogeneity of variances, then a permutational analysis of variance (PERMANOVA) test was run using the same model. PERMANOVA analyses were based on a Euclidean dissimilarity matrix using a minimum of 9999 permutations of residuals under a reduced model. R (version 3.0.2; R Development Core Team 2013) was used for all statistical analysis and graphs were produced using the 'ggplot2' add-on package (Wickham 2009). Results Seawater salinity tolerance trials Survival and Growth Survival was strongly reduced at the highest salinity (Table 1). At 54 psu, only 31% of the plants were surviving by 6 weeks, and all showed signs of salt damage (blackened patches on Page

103 leaves and dead roots). At the intermediate salinity (46 psu), most plants (67%) were alive, but also showing signs of salt damage (Fig.7). Fig 7. Salt scorch on leaves of Posidonia australis after exposure to raised salinity. Root growth was significantly inhibited by Treatment (p <0.003, Table 1), with almost no growth at the highest salinity. At the intermediate salinity, growth rates at 6 weeks were less than 50% of controls (Fig. 8). Leaf growth showed no significant responses to salinity treatments but was significantly inhibited by Time, indicating a transplant effect. Table 1. Survival of seagrass Posidonia australis following 6 weeks exposure to raised salinities (46 & 54 psu) compared to seawater controls (37 psu). Salinity Treatment (psu) Survival (% ± se) ± ± ± Page

104 Fig. 8. Leaf and root growth of Posidonia australis in seawater controls (37 psu) compared with raised salinities (46 & 54 psu). Photosynthesis Rapid Light Curve parameters showed significant reductions for Treatment and Time (p <0.01) (Figs 9, 10, Table I). Maximum electron transport rates (ETRmax, p <0.001) and saturating irradiance (Ek, p <0.0001) were significantly reduced at the highest salinity (54 psu) by 4 weeks. At the intermediate salinity 46 psu, ETRmax and Ek were maintained at similar values to controls for 4 weeks but were significantly lower than controls by 6 weeks (Fig. 9 a,b). There were no changes in photosynthetic efficiency (slopes of the rapid light curves in the light limited region, α) at raised salinities compared with controls (Fig. 10). Page

105 Page

106 Fig. 10. Photosynthetic parameters of seagrass Posidonia australis (a) slope of rapid light curves in the light limited region, α, (b) maximum electron transport rate (ETR max), (c) saturation irradiance (Ek) at 0, 2, 4 & 6 weeks exposure to seawater controls (37 psu) and raised salinities (46 and 54 psu). Values are means ± SE, n = 12. Page

107 Water relations After exposure to increased salinity treatments, P. australis responded to the more negative external water potentials with more negative leaf water potentials, achieved by lowering osmotic potentials while maintaining slightly positive turgor pressures. As the salinity increased, the mean water potential of the external medium was reduced from approx MPa in ambient seawater (37 psu) to -3.2 MPa in the 46 psu treatment and 3.8 MPa in 54 psu. Water potentials were significantly more negative (p=0.001) after 6 weeks exposure in both salinity treatments compared with controls. There was a significant pair-wise decrease in water potential after 6 weeks from 37 to 46 psu, and from 46 to 54 psu (Fig. 11a). There were no significant changes in osmotic pressure between controls and the intermediate salinity. Osmotic potentials were significantly more negative only after 6 weeks at the highest salinity treatment compared with controls (Fig. 11b). Turgor pressures, derived from the balance of water and osmotic potentials, showed only minor changes from MPa in controls, whereas mean values remained similar in salinity treatments ( MPa) (Fig.11c). Water contents showed significant time (p=0.02) and treatment (p=0.02) responses after six weeks at intermediate and highest salinities (Fig.11d). Pair-wise comparisons showed significant decreases in water content between 37 psu and 46 psu, and between 46 and 54 psu. Page

108 Fig. 11. (a) water potential, (b) osmotic potential (c) turgor pressure (d) leaf water content, measured for Posidonia australis leaves at 0, 4 and 6 weeks exposure to seawater controls (37 psu) and raised salinities (46 & 54 psu). Osmolyte concentrations Ions Leaf potassium (K + ) and calcium (Ca ++ ), and rhizome (Na + ) concentrations significantly decreased as salinity increased. At the highest salinity, there were significant decreases in concentrations of K (p<0.0002) and Ca (p<0.03). Sodium and chloride (Cl - ) did not change with increased salinity (Fig. 12a). The decreases in K and Ca resulted in significant decreases in the molar ratios of Ca:Na and K:Na (Fig. 13b). Page

109 Compatible solutes - Sugars The concentration of sucrose in leaf tissue significantly increased with salinity (p<0.001), but there were no significant changes for glucose or fructose (Fig. 12c). The concentration of fructose, and sucrose in rhizomes was not significantly different between treatments and there was no glucose detected. Fig. 12 (a) Concentrations of ions and soluble sugars in leaf tissue of seagrass Posidonia australis following 6 weeks exposure to seawater controls (37 psu) and raised salinities (46 and 54 psu): calcium (Ca), chloride (Cl), potassium (K), sodium (Na). (b) molar ratios of ions, Ca:Na, K:Na, (c) concentrations of soluble sugars (sucrose, glucose, fructose), n=27. In the rhizomes, only Na showed a significant increase at the highest salinity from ca 60 at 37, 70 at 46 and 110 at 54 psu. In contrast to leaves, Ca and K remained constant with increased salinity. Mean values for ion concentrations in rhizomes were similar to those in leaves. Page

110 Amino acids In leaf tissue, several amino acids showed significant increases after 6 weeks exposure to salinity treatments (ARG, HIS/GLN, ILE, LEU, LYS ALA, PHE, PRO, SER/ASN, THR, VAL, abbreviations listed in figure legend, Fig. 13a,b). In rhizome tissue, concentrations of several amino acids showed significant increases after 6 weeks exposure to salinity treatments (ALA, GABA, GLY, HIS/GLN, PHE, PRO, abbreviations listed in figure legend, Fig. 13c). Fig. 13. Concentrations of free amino acids in seagrass Posidonia australis after 6 weeks exposure to raised salinities (46 and 54 psu) compared with seawater controls (37 psu) (n=80). (a) amino acids with concentrations <0.3 µmol g -1 DW in leaf tissue: argentine ARG, histidine/glutamine HIS/GLN, isoleucine ILE, leucine LEU, lysine LYS (b) amino acids with concentrations <1.0 µmol g -1 DW in leaf tissue: alanine ALA, phenylalanine PHE, proline PRO, serine/asparagine SER/ASN, threonine THR, valine VAL. (c) rhizomes: alanine ALA, g -amino butyric acid GABA, glycine GLY, histidine/glutamine HIS/GLN, PHE phenylalanine, proline PRO. Page

111 Summary of findings Elevated seawater salinity exposure The ecophysiology response of Posidonia australis to elevated seawater salinity revealed a very high tolerance to raised salinity, even with rapid increase in concentration without a long period of acclimation. Leaf growth was not significantly influenced by elevated seawater salinity, however a number of stress responses were apparent: root growth was reduced after 6 weeks at both the intermediate (46 psu) and highest salinity (54 psu), photosynthesis (ETR) was significantly reduced over the 2-4 week interval after exposure to 54 psu. Osmotic and water potentials were increasingly more negative with increasing salinity, but at 54 psu, these potentials could not be maintained as photosynthesis became less efficient and osmoregulation began to break down, ratios of ions changed at higher salinity: Na and Cl did not change but concentrations of K & Ca were lower, and compatible solutes increased over the trial period at raised salinity, particularly sucrose and some amino acids in the leaves. Concentrations were higher to start with in rhizomes in line with their storage function. Page

112 3.4 Mature seagrass brine exposure The effects of desalination brine (54 psu) on seagrass shoots is investigated to compare effects of brine exposure relative to seawater salinity. Tank system and experimental design Seagrass plants in culture were set up in large aquaria at indoor facilities at UWA to provide constant background conditions; temperature, light, day length. Collection and culture techniques with Posidonia australis shoots followed the methods described above for elevated seawater salinity testing. Test plants were kept in seawater under culture conditions for 5 days before commencing the experiment. Brine treatments were then imposed without an acclimation period, in order to simulate sudden exposure to brine wastewater. The experimental design is shown in Fig. 14, with 4 tanks per treatment. Natural seawater was used for Controls with a salinity of 37 psu (range 36.4 to 36.8 psu), which is equivalent to salinity in the area where plants were collected. Brine with a salinity of 54 psu was collected from the Southern Seawater Desalination Plant (SSDP) at Binningup, Western Australia at a point directly before entering the ocean discharge pipeline. Dilutions of 25% and 50% brine were made using natural seawater. SW 37 psu Experimental design 25% 42 psu 1 50% 47 psu 1 100% 56 psu 1 SW 37 psu 25% 42 psu 2 50% 47 psu 2 100% 56 psu 2 SW 37 psu 3 25% 42 psu 3 50% 47 psu 3 100% 56 psu 3 SW 37 psu 4 25% 42 psu 4 50% 47 psu 4 100% 56 psu 4 5 plants Each plant with 2 shoots Fig 14. Conceptual diagram of experimental design for brine exposure to Posidonia australis plants. Four tanks per treatment were used for experiments, with 2 weeks exposure to brine concentrations and seawater controls. Page

113 Survival, growth, photosynthesis, water relations, concentrations of ions, amino acids, and soluble sugars were measured at the end of the 2 week exposure to compare effects of brine treatments against seawater controls, following methods described for elevated seawater salinity testing. Root growth was not measured, as the 2 week long treatment time was too short to allow new roots to develop and grow. Data analyses All studied parameters tested the null hypothesis that there were not significant differences among brine treatments or sampling times. Analyses consisted of nested analysis of variance (ANOVA) design with Treatment as a fixed factor, while Tank was treated as random factor nested within Treatment. Data was transformed (log(x+1)) when necessary and ANOVA assumptions were tested. If data presented significant deviations from normality and homogeneity of variances, then a permutational analysis of variance (PERMANOVA) test was run using the same model. PERMANOVA analyses were based on a Euclidean dissimilarity matrix using a minimum of 9999 permutations of residuals under a reduced model. Results Growth in brine treatments Leaf growth was significantly affected by Treatment (p <0.001, Fig. 15). There was no difference in leaf growth between seawater controls and 25 % brine. At 50% brine (salinity 46 psu), growth rates were 25% less than controls (pairwise comparison p <0.03). There was almost no leaf growth at 100% brine (salinity 54 psu). Fig. 15. Leaf growth of seagrass Posidonia australis with exposure to 25, 50 & 100% brine compared to seawater controls. Page

114 Photosynthesis chlorophyll fluorescence There was no difference in maximum electron transport rate ETRmax between 25% brine and controls, but at 50 % brine ETRmax was significantly higher than controls (p <0.03), suggesting a stimulation of photosynthesis at a higher brine concentration (Fig. 16). Pairwise comparisons showed that maximum electron transport rates (ETRmax,) were significantly reduced after exposure 100% brine compared to seawater controls (p <0.03). (Note that ETRmax in salinity trials psu showed no difference to seawater controls following 2 weeks exposure to 46 psu but was reduced after 4 and 6 weeks.) Rapid Light Curve parameters ETRmax and alpha showed significant treatment effects but there was no significant effect for saturating irradiance (Ek). (p <0.05) Fig.17). Fig. 16. Rapid Light Curves (ETR max) (a) Initial, and (b) Final, after 2 weeks exposure to 25, 50 & 100% brine compared to seawater controls. Page

115 a b c Fig. 17. Rapid light curve (RLC) parameters of seagrass Posidonia australis with exposure to 25, 50 & 100% brine compared to seawater controls (a) maximum electron transport rate (ETR max), (b) saturation irradiance (Ek) (c) alpha, slopes of rapid light curves in the light limited region. Values are means ± SE, n = 12. Water relations There was a significant decrease in water content (approx. 10%) following two weeks exposure to 50% and 100% brine compared to seawater controls (pairwise comparison p = 0.03, 0.01 resp.) but not to 25% brine (Fig. 18a). There was a significant pair-wise decrease in water potential between 25 and 50% brine (p<0.0001), and 50 to 100% brine (p<0.0001) (Fig. 18a). There were no significant changes in osmotic pressure between controls and 25% brine but osmotic potentials became significantly more negative at 50% and 100% compared to seawater controls (Fig. 18b, Table 2). Mean water potential of the external medium was reduced with increasing brine concentrations from approx MPa in ambient seawater (37 psu) to -3.2 MPa in the 50% brine treatment and 3.8 MPa in 100% brine. Turgor pressures remained positive and showed a small but significant loss with exposure to the most concentrated brine (p<0.03 for 100% brine compared to seawater control) but not to 25 and 50 % brine (Fig.18b). Page

116 (a) Fig. 18. (a) Leaf water content, (b) water potential, osmotic potential, and turgor pressure, measured for Posidonia australis leaves following two week exposure to 25, 50 & 100% brine compared to seawater controls. (b) Page

117 Osmolyte concentrations: ions, soluble sugars and amino acids Leaf ions There were significant treatment effects (p=0.01) for leaf sodium (Na) and chloride (Cl) ions to increasing concentrations of desalination brine (25%, 50%, 100%, corresponding to salinities of 42, 46 and 54 psu) compared to seawater controls (37 psu) (Table 2). Leaf chloride ions increased in 50% and 100% brine (63 and 55 mg g -1 DW, resp.) compared to controls (49 mg g -1 DW). Potassium ions (K) showed a small non-significant decrease in 100% brine, compared to controls, from 30 to 22 mg K -1 g DW. There were small non-significant increases in calcium (Ca) ions in 25% and 50% brine but no change in 100% brine (Fig.19a). There were no significant changes with treatments in the mean molar ratios of ions, Ca:Na ( ) and K:Na ( ) (Fig. 19b). Fig. 19 (a) Concentrations of ions: calcium (Ca ++ ), chloride (Cl - ), potassium (K + ) and sodium (Na + ) in leaf tissue of seagrass Posidonia australis following 2 weeks exposure to increasing concentrations of desalination brine (25%, 50%, 100%, corresponding to salinities of 42, 46 and 54 psu) compared to seawater controls (37 psu), n=4. (b) Molar ratios of ions in leaf tissue. Page

118 Table 2. Concentrations of ions (means ± sd): potassium (K+), sodium (Na+), calcium (Ca++) and Chloride (Cl-) in leaf tissue of seagrass Posidonia australis following 2 weeks exposure to increasing concentrations of desalination brine (25%, 50%, 100%), corresponding to salinities of 42, 46 and 54 psu) compared to seawater controls (37 psu), n=4. Leaf mg K -1 g DW mg Na g -1 DW mg Ca g -1 DW SW 0% 29.83± ± ± % 26.52± ± ± % 28.63± ± ± % 22.02± ± ±10.67 mg Cl g -1 DW 48.63± ± ± ±3.88 Rhizome ions There were significant treatment effects for chloride (Cl) and sodium (Na) in rhizome tissue to increasing concentrations of desalination brine. 25%, 50%, 100%, corresponding to salinities of 42, 46 and 54 psu) compared to seawater controls (37 psu). Mean concentrations of chloride ions increased significantly from 76 mg g -1 DW in seawater controls to 110 mg g -1 DW in 100% brine (p<0.001). These values were approximately double those of chloride concentrations in leaf tissue, whereas the other ions had similar values in leaf and rhizome tissue. Sodium ions showed minor but significant decreases (p<0.002) with increasing brine concentration in pairwise comparisons (Fig. 20a). Page

119 (a) (b) Fig. 20 (a) Concentrations of ions: calcium (Ca ++ ), chloride (Cl - ), potassium (K + ) and sodium (Na + ) in rhizome tissue of seagrass Posidonia australis following 2 weeks exposure to increasing concentrations of desalination brine (25%, 50%, 100%, corresponding to salinities of 42, 46 and 54 psu) compared to seawater controls (37 psu), n=4. (b) Molar ratios of ions. Compatible solutes: Soluble sugars The concentration of sucrose in leaf tissue increased significantly with increasing concentrations of brine, from 11.9 to 15.7 mg g -1 DW (25 to 100% brine) compared to seawater controls, 7.7 mg g -1 DW (Table 3). Concentrations of glucose and fructose were very low (Table 3), often below detectable limits and there were no significant changes with treatments (Fig. 21). An unknown sugar which was possibly galactose, was also detected. Page

120 Table 3. Mean concentrations of soluble sugars in leaf tissue following 2 weeks exposure to increasing concentrations of desalination brine (25%, 50%, 100%, corresponding to salinities of 42, 46 and 54 psu) compared to seawater controls (37 psu), n=4. Leaf Sucrose mg g -1 DW Glucose mg g - 1 DW Fructose mg g -1 DW SW 7.65± % 11.89± % 12.87± % 15.70± Fig.21 Concentration of soluble sugars in (a) leaf tissue (sucrose and an and unknown sugar, possibly galactose), (b) leaf tissue (fructose, glucose), (c) rhizome tissue (sucrose and an and unknown sugar, possibly galactose) of seagrass Posidonia australis following 2 weeks exposure to increasing concentrations of desalination brine (25%, 50%, 100%, corresponding to salinities of 42, 46 and 54 psu) compared to seawater controls (37 psu), n=4. Page

121 Amino acids In leaf tissue, only two amino acids were significantly increased with brine exposure, proline (p= 0.01) and tyrosine (p= 0.008, Fig. 22a). (a) (b) (c) Fig. 22. Concentration of free amino acids in seagrass leaves Posidonia australis following exposure to brines (25, %) compared to seawater controls (37 psu). (a) only proline PRO and tyrosine TRY increases were significant. (b) all amino acids with concentrations <5 µmol g -1 DW in leaf tissue: argentine: alanine ALA, gamma-amino butyric acid GABA, histidine/glutamine HIS/GLN, ammonium NH 3, proline PRO, serine/asparagine SER/ASN, valine VAL. (c) all amino acids with concentrations <0.3 µmol g - 1 DW in leaf tissue: argenine ARG, isoleucine ILE, leucine LEU, phenylalanine PHE, threonine THR, tyrosine TYR. (d) all amino acids with concentrations <0.15 µmol g -1 DW in leaf tissue: asparagines ASP, cysteine CYS, glutamine GLU, glycine GLY, lysine LYS, methionine MET. (d) Page

122 In rhizome tissue, concentrations of all amino acids tested were significantly increased with exposure to brines. Values were lower for 100% brine than 25 and 50 % brine. (Fig. 23a,b). Fig. 23. Concentration of free amino acids in seagrass rhizome Posidonia australis showing significant increases following 2 weeks exposure to brines (25, %) compared to seawater controls (37 psu). (a) amino acids with concentrations <10 µmol g -1 DW in rhizome tissue: alanine ALA, argentine ARG, gamma-amino butyric acid GABA, histidine/glutamine HIS/GLN, isoleucine ILE, proline PRO, serine/asparagine SER/ASN, threonine THR, valine VAL. (b) amino acids with concentrations <0.6 µmol g -1 DW in rhizome tissue: ASP, cysteine CYS, glutamine GLU, glycine GLY, leucine LEU, lysine LYS, methionine MET, phenylalanine PHE,, tyrosine TYR. Page

123 Summary of findings Brine exposure to mature plants P. australis exhibits alteration in morphological characteristics and selected physiological processes: leaf growth was reduced at both intermediate (50%) and highest brine (100%) concentrations water content and water potential decreased in leaves with brine exposure of 50% and above turgor pressure dropped in leaves at 100% brine exposure ratios of ions changed at higher brine exposure: Cl increased in both leaves and rhizomes sucrose in leaves increased markedly with increase in brine concentration, and amino acids in rhizomes increased with brine exposure. Page

124 3.5 Seedling brine exposure This test investigated the effect of increased brine concentration exposure to seagrass Posidonia australis seedlings over 7 weeks, followed by 3 weeks recovery in seawater. Methods (seedlings) Fruit of Posidonia australis were collected from Woodman Pt. and maintained in ambient seawater (23 C) to allow natural dehiscence to occur over 3 days in a 350-L holding tank, receiving natural light and through-flow filtered seawater (15 L/min). Dehiscence of mature fruit resulted in seeds with a mean seed length of 15 mm (±1 mm) falling to the bottom of the holding tank. Intact seeds of P. australis with the first stages of a developing shoot and radical in evidence were collected within 1 day of release from the fruit and placed in 10 L plastic containers with approx. 10 cm layer of unsorted silica sand with addition of organic matter (1.5% by weight as described in Statton et al. 2013) and approx. 8 L of seawater to provide a water depth of 30 cm. Ten seeds were placed on the sediment surface of each container. Treatments were 25 % brine and 50 % brine diluted with seawater, and100% brine, compared seawater controls (37 psu) with four replicates per treatment. Conditions were held constant in a controlled environment facility, with air temperature of 22 o C, and 14 hours light and10 hours darkness, corresponding to the mid-summer light regime near Perth at latitude 33 o S. Lighting was provided by overhead LED lamps with irradiance of 500 µmol m -2 s -1 at the water surface. The water was aerated with compressed air delivered through one bubble stone per container. Salinities were checked daily and held constant by adding de-ionised water. Seeds were held for 7 weeks in treatments and then returned to ambient seawater over 4 days and held for a further 2 weeks to test for recovery. Ten seedlings were harvested from each container 10 weeks after planting. Seedlings were photographed in their fresh state then separated into their component parts (leaf shoot, roots, seed). All material was then dried in an oven at 60 C for 72 hours and weighed. Total root length (mm) per seedling was measured from digital images using ImageJ software. Results - Seedlings Leaf shoot and root biomass showed a decrease with increasing brine concentrations, which was statistically significant in 50% and 100% brine (Fig.24).Total root length, measured as a Page

125 more sensitive indicator of root growth response, was significantly lower in 50% and 100% brine than both seawater controls and 25% brine (Fig.25). Seedlings were very resilient to a long exposure to brine. In 100% brine, there was almost no leaf growth and no root growth for 7 weeks but growth began within a week of exposure to ambient seawater (Fig. 26). More leaf growth occurred at 50% brine concentration but root growth was inhibited. At 25% brine, there was almost no difference with seawater controls (Fig. 26). The main response was inhibition of root growth, whereas some shoot growth was less affected. The leaf sheath has been shown to protect expanding young leaves of Posidonia australis (Tyerman et al. 1989) and it is likely that this also occurs in the first stages of the life cycle in germinating seedlings. Fig. 24. Leaf shoot and root biomass per seedling for seagrass Posidonia australis showing significant decreases with 7 weeks exposure to brines (25, 50 & 100%) followed by three weeks recovery in seawater (37 psu). Fig.25. Total root length per seedling for seagrass Posidonia australis produced during 7 weeks exposure to brines (25, 50 & 100%), followed by three weeks recovery in seawater SW (37 psu). Page

126 (a) Seawater control (b) 25% brine (c) 50% brine (d) 100% brine Fig. 26. Posidonia australis seedlings after 7 weeks exposure to brine treatments followed by 3 weeks recovery in ambient seawater. Seedlings from 50 % brine had reduced leaf shoot and root growth compared to seawater controls. Seedlings from 100% brine had no growth but after 3 weeks in seawater, there was some recovery with pale stunted shoot growth, although most seedlings had not begun to produce roots. Summary of findings Brine exposure to seedlings P. australis seedlings exhibit a high sensitivity to desalination brine exposure: leaf shoot and root biomass decreased markedly with increase in brine concentration, total root length decreased markedly with increase in brine concentration evidence of recovery in seedling growth with cessation of brine exposure Page

127 3.6 Summary of response to elevated seawater hypersalinity and brine Physiological endpoints for sublethal response to salinity (Tables 3,4 &5). Water relations osmotic potential and water potentials, turgor pressure and water content of leaves showed significant responses to counter the effects of salt-induced dehydration. Each contributes to the overall mechanism of the plant to counter stress from raised salinity. Osmolytes concentrations of ions, soluble sugars and amino acids. Ions potassium (K) and calcium (Ca) concentrations in the leaves decreased after 6 weeks at the highest salinity (54 psu) but in the shorter brine trials lasting 2 weeks, there was no change, indicating that a measurable response would require longer exposure. Soluble sugars - sucrose showed a strong response, dependent on salinity level and exposure time, e.g. x2 increase after 2 weeks in 100% brine and x4 increase after 6 weeks at 54 psu seawater. Other sugars (fructose, glucose) showed only minor or no response. Amino acids most of the amino acids measured in leaf tissue showed increases after 6 weeks exposure to raised salinity but only proline and tryrosine showed statistically significant increases after 2 weeks exposure to 100% brine. These increases suggest that amino acid composition is not stable under increasing hypersalinity exposures and are likely acting on membrane permeability and ion transport, proline in particular potentially facilitating in salt transport from the roots to the shoot (Rai 2002). Chlorophyll fluorescence - parameters showed delayed onset of response, such as reduced electron transport rate (ETRmax). Results were significant in relation to the values decreasing with increased hypersalinity, but highly variable for individual leaves requiring measurement of multiple leaves per individual plant. Page

128 Table 3. Salinity trials with exposure to 2 levels of raised salinity (46, 54 psu) for 6 weeks corresponding to the salinity of 50% and 100% desalination brine. NS=not statistically significant. Measured endpoint Method Response after 4 &6 weeks compared to SW controls Salinity at which response was significant (p<0.05) Time (weeks) Growth Leaf marking transplants Leaf - no treatment effect Roots reduced growth Photosynthesis Chlorophyll fluorescence Reduced ETRmax Water relations Water content Reduced water content 54 6 Water relations Osmotic potentials Water potential Reduction (more ve), results variable Reduction (more ve), results variable (46 NS) 54 (NS -) Turgor pressure loss of turgor, results variable 46 4 Ions Na, Cl, Ca, K K sig decrease 54 6* *after 6 weeks Ca sig decrease 54 6* Cl high, Na low, no sig change Sugars* *after 6 weeks Sucrose sig. increase in leaf, not rhizome 46 6* glucose below detect., fructose no change Amino acids *after 6 weeks Sig increase in most leaf amino acids (ALA, ARG, GABA, HIS/GLN, ILE, LEU, LYS, PHET, PRO, Ser/ASN, THR, VAL * 6* Sig increase in rhizome amino acids ALA, ARG, ASP, GLY, HIS/GLN Survival Shoot death 67% still alive 31% still alive Page

129 Table 4. Brine trials -exposure to brines for 2 weeks, 3 levels of desalination brine, 25%, 50% and 100%. Measured endpoint Method Response after 2 weeks compared to seawater controls % brine at which response was significant (p<0.05) Growth (mature plants Leaf marking Root growth at harvest 25% 50%, no Same as leaf growth >25%, <50% >25%, <50% Photosynthesis Chlorophyll fluorescence Reduced ETRmax sig in 100%, cp to 50%, 25% & 0%. Ek no diff. >50% Reduced Alpha sig. in 100%, cp to 50%, 25% & 0%. >50% F0/FM sig reduced in 50%, 100%, cp to 25% & 0%. >25%, <50% Water relations Water content Reduced water content 50%, 100% brine, not 25% >25%, <50% Water relations Osmotic potentials Reduction (more ve) water potential, 50 & 100% brine. >25%, <50% Water potentials Reduction (more ve) sig only in 100%, >50%, <100% Turgor pressure loss of turgor borderline sig@100% =50%, <100% Ions Na, Cl, Ca, K Small increase in Ca up to 50%, then small decrease in 100%, no sig change in Na, K. Cl small increase in 50, 100% 25-50% Compatible solutes Soluble sugars Sig. increase in sucrose (x2@100%), glucose & fructose borderline or below detection limits >50% Amino acids Sig increase in leaf proline 100%, tyrosine (TYR) in % Sig increase in most rhizome amino acids <25% Page

130 Table 5. Seeds and germinating seedlings exposure to brines for 7 weeks, followed by recovery in seawater for 2 weeks. Measured endpoint Method Response after 2 weeks compared to seawater controls % brine at which response was significant (p<0.05) Growth (seedlings) over 7 week, followed by 2 weeks recovery in seawater control Leaf growth Root growth Biomass increase 7 brine (54-56 psu) no leaf or root growth, followed by recovery with shoot 37 psu over 2 weeks 7 brine (46-48 psu) some leaf but no root growth, followed by recovery with shoot growth@ 37 psu over 2 weeks >25% 7 brine (42 psu) leaf and root growth similar to seawater controls followed by recovery with shoot growth@ 37 psu over 2 weeks Conclusion Seedlings were resilient to brine exposure but showed distinct stress response in leaf and root growth. In 100% brine, almost no leaf growth and no root growth was evident for 7 weeks; after being placed in seawater at ambient seawater salinity, shoot growth began within a week. The leaves, however, were pale and stunted, indicating that much of the storage reserve (starch and nutrients) in the seeds had been consumed in sustaining the seedlings during the exposure to brine with a very high salinity. More leaf growth was evident at 50% brine concentration but root growth was still markedly inhibited. At 25% brine, there was almost no difference in leaf and root growth compared to seawater controls. Root material appears to more sensitive to hyperslinity evidenced by the inhibition of root growth, whereas shoot growth was less affected. The leaf sheath has been shown to protect expanding young leaves of Posidonia australis (Tyerman et al. 1989) and it is likely that this also occurs in germinating seedlings. Seedlings of P. australis are only available for a very short window of time, approx. 2 weeks in early December when the seeds are shed. They exhibit distinct responses to brine exposure, are relatively easy to culture in controlled conditions and can be assessed in large numbers. They would therefore provide a useful early life cycle bioassay to assess the effects of salinity wastewater discharge under laboratory conditions. Page

131 3.7 References Kramer, P. J. and J. S. Boyer (1995). Water Relations of Plants and Soils. Academic Press, San Diego, USA. Pp Marín-Guirao, L., Sandoval-Gil, JM., Bernardeau-Esteller, J, Ruíz, JM, Sánchez-Lizaso, JL (2013). Responses of the Mediterranean seagrass Posidonia oceanica to hypersaline stress duration and recovery. Marine Environmental Research 84: McMahon, K., Collier, C. J., Lavery, P.S. (2013). Identifying robust bioindicators of light stress in seagrasses: A meta-analysis. Ecological Indicators 30: Orth, R.J., Carruthers, T. J. B.,Dennison, W. C., Duarte, C.M., Fourqurean, J.W., Heck, Jr., K. L., Randall Hughes, A., Kendrick, G. A., Kenworthy, W. J.,Olyarnik, S., Short, F. T., Waycott, M., Williams, S. L A global crisis for seagrass ecosystems. Bioscience 56: Pulich WM (1986) Variations in leaf soluble amino acids and ammonium content in subtropical seagrasses related to salinity stress. Plant Physiol 80: R Development Core Team (2013) R: A language and environment for statistical computing. R foundation for Statistical Computing, Vienna, Austria. Rai, V.K Role of amino acids in plant responses to stressors. Biologia Plantarum 45(4) Ralph, PJ., Gademann R Rapid light curves: a powerful tool to assess photosynthetic activity. Aquatic Botany 82: Ruiz, J.M., L. Marın-Guirao, J.M. Sandoval-Gil. (2009) Responses of the Mediterranean seagrass Posidonia oceanica to in situ salinity increase. Botanica Marina Sandoval-Gil, J., Marín-Guirao, L., Ruíz, JM (2012a). Tolerance of Mediterranean seagrasses (Posidonia oceanica and Cymodocea nodosa) to hypersaline stress: water relations and osmolyte concentrations. Marine Biology 159: Sandoval-Gil, J. M., Marín-Guirao, L., Ruíz, JM (2012b). The effect of salinity increase on the photosynthesis, growth and survival of the Mediterranean seagrass Cymodocea nodosa. Estuarine, Coastal and Shelf Science 115: Page

132 Schwarz, A.-M., M. Björk, Buluda, T., Mtolera, M., Beer, S. (2000). Photosynthetic utilisation of carbon and light by two tropical seagrass species as measured in situ. Marine Biology 137: Statton, J., Cambridge, M.L., Dixon, K.W., Kendrick, G.A Aquaculture of Posidonia australis seedlings for seagrass restoration programs: effect of sediment type and organic enrichment on growth. Restoration Ecology. 21: Statton, J., Kendrick, G.A., Dixon, K.W., Cambridge, M.L Inorganic nutrient supplements constrain restoration potential of seedlings of the seagrass, Posidonia australis. Restoration Ecology 22: Short, F.T, Duarte, C.M Methods for the measurement of seagrass growth and production. Ch. 8. In: (eds) Short, F. T., R. G. Coles (2001). Global Seagrass Methods.. Elsevier, Amsterdam, The Netherlands. Pp Slimestad, R. (2006). Thermal stability of glucoser and other sugar aldoses in normal phase high performance liquid chromatography. J. Chrom. A 1118: Tyerman SD (1982) Stationary volumetric elastic modulus and osmotic pressure of the leaf cells of Halophila ovalis, Zostera capricorni and Posidonia australis. Plant Physiol 69: Tyerman, S., Hatcher, A., West, RJ, Larkum, AWD. (1984). Posidonia australis growing in altered salinities: Leaf growth, regulation of turgor and the development of osmotic gradients. Functional Plant Biology 11: Tyerman SD, Hatcher AI, West RJ, Larkum AWD (1984). Posidonia australis growing in altered salinities: leaf growth, regulation of turgor and the development of osmotic gradients. Aust J Plant Physiol 11: Tyerman SD (1989) Solute and water relations of seagrasses. In: Larkum AWD, Mc Comb AJ, Sheperd SA (eds) Biology of seagrasses: a treatise on the biology of seagrasses with special reference to the Australian Region. Elsevier, Amsterdam, pp Warren, C. R., and Adams, M.A. (2000). Capillary electrophoresis for the determination of major amino acids and sugars in foliage: application to the nitrogen nutrition of schlerophyllous species. Journal of Experimental Botany 51(347): Page

133 4. Biomarker development - Bivalves Morphological and physiological effect of elevated seawater salinities and desalination-derived brine on key bivalve species. Julie Mondon 1, Shaun Davis 1, Elizabeth Cain 1 1. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment Deakin University Warrnambool, Victoria Australia Key questions 1. To what extent does elevated salinity and desalination-derived brine exposure result in a deleterious stress response in bivalves? 2. Which response, or suite of responses, are most suitable as biomarkers of effect from hypersalinity? 4.1 Overview Bivalves were investigated as model organisms to identify representative salinity threshold ranges of common mollusc species, enabling identification of: Dose-response thresholds Suitability for replication and deployment at distance from discharge, and Model testing procedures for future studies 4.2 Pilot investigation Hypersalinity stress response - Xenostrobus pulex Qualitative / quantitative measures of stress summary Indicators of whole organism stress Transferable biomarker from laboratory to in situ exposure Relatively low cost, easy to perform & replicate, minimal technical expertise Adaptable to other species Pilot organism - Xenostrobus pulex (Little black horse mussel) Locally available in abundant mat-forming dense populations Wide geographical distribution across southern Australia; from southern WA, across to Victoria, around Tasmania and up to northern New South Wales Small in size and robust to handling; length up to 2.5 cm Intertidal species Page

134 Figure 1. X.pulex (Lamarck, 1819) Xenostrobus pulex (shell length 13.5 mm ± 1.5) were collected from the low intertidal zone at Warrnambool Harbour ( S, E). Epiphytic growth on the shell was removed prior to transportation in 200 µm filtered seawater to the Deakin University laboratory. Animals were divided into holding aquaria, ten individuals per 8 L glass tank containing 4 L of filtered aerated natural seawater, and left to depurate without feeding for three days (Figure 2). Following depuration, a feeding regime comprising dried Spirulina (Life Stream) was provided every 3 days; feed volume was governed by the clearing rate (ie. uptake of Spirulina from the water column over 12 hours). Tanks were housed in an air conditioned external ambient air temperature of 18 C, and maintained under a natural photoperiod. A 50% water exchange occurred every second day. Water quality parameters were measured daily and maintained at 17.3 C ± 1.39, ph 8.05 ±0.28, 32 ± 1.57 and DO 8.96 mg/l ± All seawater for husbandry and experimentation was collected from Warrnambool harbour and primary filtered to 20 µm on site prior to transportation to Deakin University. Secondary filtration (4 µm canister filter) occurred at Deakin prior to final storage in sealed aerated storage containers at 17.4 C ± Salinity concentration for test procedures used filtered seawater modified by the addition of Ocean Pure synthetic sea salt; incremental steps of 10 g Ocean Pure per L filtered seawater were used to raise seawater salinity by 1 salinity unit (psu) to the required salinity. Modified salinity seawater was held overnight prior use in treatment tanks. Pilot seawater salinity exposure test summary 10 day laboratory acclimation period temperature and salinity controlled conditions ambient daylight photoperiod filtered natural seawater / filtered natural seawater + artificial sea salt 7 day exposure test Page

135 B. Cain Figure 2. Test aquaria holding mussels in aerated tanks under ambient light. Test treatments Sudden salinity increase (pulse) exposure versus gradual salinity increase (ramped) exposure was investigated. Seven 8 litre tanks per salinity exposure regime (sudden and gradual) were established, with seven treatments maintained at the following partial salinity units (psu); 35, 45, 55, 65, 75, plus 32 (control), and 32 (procedural control, comprising 50% natural seawater mixed with 50% Ocean Pure and deionised water). The control salinity value represents in situ salinity from where X. pulex and seawater was collected. Sudden pulse increase in salinity was generated using filtered seawater and artificial sea salt, which was maintained at the required salinity for the duration of the test procedure by 50% volume water changes daily. Gradual salinity increase (ramping) was achieved by an increase of 1.5 psu, 2.6 psu, 4.6 psu, 6.5 psu, 8.6 psu respectively per day. Fifty % volume water changes were conducted daily with elevated salinity seawater to reach the desired final salinity. Throughout the pulse and ramped exposures mussels were fed every three days as described above. Water quality parameters were measured daily and maintained at C ± 0.38, ph 8.26 ±0.04, DO 10.4 ±.07 mg/l. Oxygen consumption trials commenced when elevated salinity exposure reached the designated concentration. The test trails were replicated three times (Table 1). Table 1. Experimental design for replicated seawater salinity 7 day exposure trials. Exposure trial SW Salinity (psu) Replicate x2 Replicate x3 Sudden Single 32 control 32 control 32 control Pulse 32 procedural control 32 procedural control procedural control Page

136 Exposure trial SW Salinity (psu) Replicate x2 Replicate x3 Gradual - Ramped 32 control 32 control 32 control 32 procedural control 32 procedural control 32 procedural control Response of X.pulex to elevated salinity exposure was measured across a suite of endpoints. Effect endpoints summary Physiological - metabolic rate (oxygen consumption) Morphological - structural / functional change - growth / size Behavioural - feeding - movement - defence response Metabolic rate - Oxygen Consumption Three individual X. pulex per treatment exposure tank were placed in separate respiration chambers to calculate oxygen uptake as a measure of metabolic rate. Four 10 ml glass vials were used as respiration chambers, each fitted with a Loligo micro sensor spot, 6 mm glass magnetic stirrer (flea), and protective stainless steel mesh suspended above the base to prevent agitation of the mussel by the spinning of the stirrer. A single fibre optic sensor cable per chamber was attached using Velcro straps containing an inbuilt sensor adapter. Chambers were calibrated daily according to Loligo OXY4-Mini transmitter calibration procedure (Loligo 2005). The linear calibration used aerated 100% saturated seawater and 0% saturated solution containing 1g sodium sulphite (AnalaR) in 100 ml seawater. Chambers were cleaned with 75 C distilled water prior to each run. All chambers were submerged in a treatment tank, with three chambers containing one mussel per chamber, and a fourth chamber left unoccupied as a Page

137 blank. Submerged chambers were left unsealed for 30 minutes to reduce handling stress, then sealed within the treatment tanks, reattached to the fibre optic sensor cables and moved into a temperature controlled water bath situated on top of a magnetic stirrer base station (Figure 3). The water bath was maintained at 17 o C using TempCtrl software (Loligo 2005) which controlled a submersible Ehiem water pump and cooling sump (Figure 4). Oxygen concentration reading output in mg/l was generated by OXY4-mini transmitter software (Figure 5). Each test period was conducted over a duration of 2 hours, at the end of which the oxygen concentration was low but not completely consumed. Mussel length, weight and volume were recorded prior to being returned to their treatment holding tank. The methodology described was repeated for all treatments and controls. For each test replicate, three individuals (treatment and controls) were sacrificed and fixed in 10% buffered formalin for histological assessment. E.Cain Figure 3. Representation of bivalves held in sealed respiration chambers containing one oxygen sensor dot per oxygen probe channel. Page

138 E. Cain Figure 4. Representation of respiration equipment set up showing glass chambers containing one oxygen sensor dot per oxygen probe channel, and cooling to maintain constant temperature during the measurement phase. Figure 5. Representative example of decline in oxygen consumption in X.pulex held in a sealed respiration chamber over a 2 hr duration. Ch1 = channel 1 (reading from blank chamber not containing a bivalve); Ch2, 3, 4 = channels 2 to 4 (readings from one bivalve per chamber). Page

139 Respiration rate (RR) RR calculation is based on Oxy4 mini transmitter values Conversion to rate values Statistical analyses Effects of treatments, exposure regime (gradual verses sudden) and duration of oxygen consumption trial, were assessed using a generalized linear model (GLM). Homogeneity of variance of the calculated values was tested using Levenes test, followed by Analysis of Variance (ANOVA). Post-hoc (Tukeys) tests were used to identify where differences occurred between salinities and mode of exposure, against controls (SPSS). Each biological replicate represents a mean value of three mussels from each treatment. All data are represented as mean ± standard error of means (SEM). Differences were considered significant if the probability of Type 1 error was equal to or less than 0.05 (p<0.05). The Effective Concentration 50% (EC50) for gradual and sudden exposures was determined by point estimate using Probit analysis. The EC50 point estimate value represents the concentration at which test individuals are likely to show a 50% reduction in oxygen uptake. Results - Oxygen Consumption Statistically significant differences between salinity treatments and oxygen uptake within the first hour of the respiration trial were evident for both gradual and sudden exposure regimes (Table 2). During the second hour no significant differences between ramped and pulse salinity increase exposure was evident. Whole animal weight corrected values represented in Figure 6 clearly indicates the change in oxygen consumption over the first hour relative to salinity exposure. Page

140 Table 2. Summary of ANOVA for gradual verses sudden salinity exposures and weight corrected verses non-weight corrected values representing oxygen uptake across a 2 hour respiration trial for X. pulex. Bold values indicate statistically significant values (p<0.05). 1 st hour Non-weight corrected Weight corrected 2 nd hour Non-weight corrected Weight corrected Gradual Ramped F6,14= p= F6,14= P= F6,14= p=0.416 F6,14= P= Sudden Single pulse F6,14= p=0.026 F6,14= P= F6,14= p=.368 F6,14= P= Figure 6. Mean ± SE X. pulex oxygen consumption rate during the first hour of respiration trials indicating significant difference between whole animal weight-corrected values and whole animal non-weight corrected values. Page

141 Significant statistical differences in mean weight-corrected oxygen consumption values were evident between salinity concentrations, but were not significantly different between gradual and sudden exposure at each measured salinity (Table 3). Table 3. Comparison between mean weight-corrected oxygen uptake in X.pulex across the exposure salinity range and rate of salinity change (ie. sudden verses gradual). Bold values indicate the highest oxygen consumption. Letters indicate where statistically significant differences occur across salinity treatments. Salinity Sudden pulse exposure O2 uptake (mg/l) Std error Significant difference between salinities P<0.05 Gradual ramped exposure O2 uptake (mg/l) Std error Significant difference between salinities P<0.05 Significant difference - sudden verses gradual B B p> A A p> B BC p> C BC p> D C p> D C p>0.05 Respiration rate (RR) Effective concentration (EC) point estimates Point estimate calculations to determine the effective concentration in oxygen uptake indicate that onset of a low level (10%) uptake reduction is likely to occur at lower mid-range concentration salinities for both sudden and gradual exposure, whereas the 50% uptake reduction estimate (EC50), after a single pulse salinity exposure is likely to be reached at a lower salinity (44 psu) compared to a gradual salinity increase (52 psu). The difference in onset of a 90% reduction in oxygen uptake between single and ramped exposure is similar (61-62 psu) (Table 4), indicating that response to the lower mid-range and higher salinity exposures is likely to be dependent on the actual salinity and independent of the rate of change in salinity, whereas the mid-range salinities are likely to elicit a lag in reduction of oxygen uptake with gradual increase in exposure. Page

142 Table 4. Effective concentration (EC) point estimates for oxygen uptake inhibition in X.pulex exposed to gradual and sudden salinity change based on weight-corrected oxygen consumption (1 hour). Oxygen uptake EC10 psu EC50 psu EC90 psu Sudden pulse Gradual - ramped Overall, oxygen consumption decreases with salinity increase, with the exception of a peak in oxygen uptake representative of hormesis (ie. low-dose stimulation) in oxygen uptake at 35 psu. Mussels exposed to 65 psu and 75 psu salinity exhibited the lowest oxygen uptake dropping to between 3-5% of the oxygen consumption exhibited by individuals at ambient 32 psu exposure. Mussels exhibiting a relatively greater tolerance under gradual verses sudden salinity change (EC50 44 psu vs. EC50 52 psu), suggests a higher degree of acclimation has occurred in response to gradual salinity increase. However, tolerance to longer term exposures beyond a 7 day exposure, as has been used in this test protocol. It is clear that exposure to higher salinities (60 psu), irrespective of being a pulse or gradual increase, is likely to result in significant decline in the respiration rate of X.pulex; effectively the coping mechanism or acclimation efficiency is reduced significantly across the range of psu upwards. Previous research has established that marine mussels can acclimate to significant salinity fluctuations (Khlebovich et al. 1973, Amende et al.1980, Bakhmet et al. 2012). Under a gradual salinity gradient the acclimation efficiency of X. pulex could be as high as 26 psu above that recorded at in situ in the field. Morphological change - Digestive gland atrophy Mussels were haphazardly selected from each treatment tank at the end of the treatment exposure period. The soft body tissue was removed by inserting an oyster knife between the valves on the bearded side close to the hinge point at the base of the valves. A scalpel was then used to separate the posterior abductor muscle from the inner shell before removal of the whole animal from the shell, and immediately fixed in 10% buffered formalin. No pre-fixation sectioning of X. pulex was required as the width of the whole soft body is <5 mm. Fixed samples were dehydrated across a graded series of ethanol from 30% to 100%, cleared in Histolene and impregnated with paraffin wax using a Reichart-Jung Histokinette processor over a 22 hour cycle. Wax impregnated samples were then embedded in paraffin wax using a Page

143 Tissue-Tek II Embedding Centre. Embedded tissues were sectioned to 5 µm using a HM 325 Microm Microtome, floated onto glass slides, and stained with standard Haemotoxylin and Eosin (H&E). Sections were viewed under an Olympus BX51 microscope fitted with a Motic10 digital image capture system. A semi-quantitative scale adapted from Lauenstein et al. (1998) was used to calculate the level of digestive gland atrophy present (Table 5). Table 5. Digestive gland atrophy scale morphological description Scale Description 0 Normal wall thickness in most tubules (0% atrophy), lumen nearly occluded, few tubules even slightly atrophied 1 Average wall thickness less than normal, but greater than one-half normal thickness, most tubules showing some atrophy, some tubules still normal 2 Wall thickness averaging half normal thickness. 3 Wall thickness less than one-half, most tubules walls significantly atrophied, some walls extremely thin (fully atrophied). 4 Wall extremely thin (100% atrophied), nearly all tubules affected. Statistical analyses ANOVA and post hoc (Tukey s) tests, as per description above, were used to identify significant differences in the semi-quantitative scale for digestive gland atrophy. Effective Concentration point estimates at 10 and 50% were calculated using Probit analysis. Results Digestive gland atrophy Significant histopathological alteration in the digestive gland of X. pulex was observed in individuals exposed to higher salinity concentrations. Reduction in digestive epithelium thickness indicates wasting or degeneration of cells and tissue structure. Narrowing of the digestive tubule wall, with corresponding increase in central lumen area, was observed in digestive glands; the degree of tissue loss increasing with higher salinity exposure (Figure 7; Figure 8). In some cases shedding of basophilic cells into the lumen was observed, resulting in a breakdown of the structural integrity of the epithelium. Digestive glands in which basophilic cell shedding occurred were difficult to categorise and were excluded from the scale, but are indicative of the epithelial cell loss movement into the lumen. In most cases significant epitheial cell loss resulted in extremely narrow digestive gland walls, with abnormally enlarged Page

144 lumen (eg. Figure 8; see 65 psu and 75 psu). The degree and rate of increase in digestive epithelial reduction is lower after a sudden salinity change, and significantly higher at 45 psu upwards after a gradual salinity change exposure (Table 6). Figure 7. Digestive gland atrophy (based on a 0-4 scale) in X. pulex exposed to gradual and sudden increase in salinity. Figure 8. Representative stages of digestive atrophy in X.pulex exposed to increasing seawater salinity exposure over 7 days. Circles indicate representative digestive tubule lumen area epithelial wall thickness. Page

145 Table 6. Comparison between mean digestive gland (DG) atrophy score in X.pulex across the exposure salinity range and rate of salinity change (ie. sudden verses gradual). Bold values indicate the highest DG score values. Letters indicate where statistically significant differences occur across salinity treatments. Salinity Sudden - pulse exposure DG Scale (0-4) Std error Significant difference between salinities P<0.05 Gradual - ramped exposure DG Scale (0-4) Std error Significant difference between salinities P<0.05 Significant difference - sudden verses gradual C D C D B C P< A C P< A B P< A A P<0.05 Analysis of change in digestive gland atrophy indicates that gradual salinity increase exposures are likely to elicit an EC50 of 42.5 psu, whilst sudden salinity increase exposures are likely to elicit a lower EC50 of 38 psu (Table 7), meaning onset of alteration occurs earlier and at a lower concentration. Table 7. Effective concentration point estimates for digestive gland atrophy above ambient salinity condition in X. pulex exposed to gradual and sudden (pulse) elevated salinity exposure. Elevated salinity exposure EC10 psu EC50 psu Sudden - pulse Gradual Histological atrophy in the digestive gland tubule of X. pulex occurs when exposed to elevated salinity. Degeneration of tubule epithelial cells hinders further digestion of nutrients due to the marked reduction in absorptive surface area in the tubule. In some cases significant shedding of basophilic cells and cellular material into the lumen also occurs; destroying the epithelial cell structure. Page

146 Digestive gland atrophy is likely to be influenced by the elevated solute concentration content in the lumen (Bœuf et al. 2001). Based on osmotic concentration, epithelial cells will lose fluid to the tubule lumen (Randall et al. 2002), thereby changing the cell s overall water balance. At the same time elevated change in salinity can physically damage the cell membrane itself, with concentrated salts altering the protein structure. Shedding of cells into the lumen has been observed in response to other contaminates such as copper (Sheir et al. 2010, Al-Subiai et al. 2011) and cadmium (Sheir et al. 2010) where it was described as necrosis of digestive tubule. Evidence from this study indicates that necrosis occurs at higher salinities. At lower to medium range salinity concentration (32-45 psu) multiple tubules with very small haemolymphatic spaces exist between tubules in the digestive gland of X. pulex. The area of haemolymphatic space increases with elevated salinity, indicating a corresponding degradation of interstitial tissue. Loss of interstitial tissue could be attributed to reduction in tissue water content at higher salinities (Shumway 1977). Haemolymphatic fluid space was more evident in sudden increase salinity exposure individuals than those exposed to gradual increase. EC50 values for digestive gland atrophy were lower than oxygen consumption trial EC50 s. A similar trend appears with sudden salinity increase exposures exhibiting a lower EC50 value than the gradual salinity increase exposures. It is unknown how longer term exposures would affect X. pulex. Oxygen uptake typically increases with digestion activity (Randall et al. 2002); greater food uptake requires more oxygen for metabolism. Degeneration of the epithelial layer of the digestive gland tubules, resulting in lower digestive uptake would require lower oxygen levels for metabolic respiration. Less oxygen is needed to generate and maintain current metabolism and energy expenditure, ie. lower metabolic energy would translate to decreased energy expenditure. Behavioural change - Semi-quantitative differential analysis Qualitative behavioural observations were made during each exposure trial. Uptake of feed was determined by the area of Spirulina settled on the base of each tank at 12 hours post feed. Movement of animals was recorded over two hours after being physically disturbed in tanks. Valve closure behaviour was recorded daily at set periods throughout the day. Special emphasis was paid to the period surrounding oxygen uptake measurement. Byssal threads were recorded daily. Particular emphasis was paid to the period following handling after oxygen uptake Page

147 measurement. Behavioural changes were scored across a gradient from none, mild, moderate or severe relative to the control. A semi-quantitative differential response scale was used to indicate the prevalence of change identified in bivalve behaviour (Table 8). Decreased feeding rate (reduced feed uptake), decreased movement after disruption such as tank cleaning, slower valve closure after disruption, and increased duration of valve closure were observed at high salinity exposures. No behavioural changes were observed in 32 psu and 35 psu treatments. Changes in behaviour were evident at exposures of 45 psu and above, and increased in prevalence with increasingly salinity. Table 8. Behavioural change activity in X. pulex with exposure to elevated salinity. Behavioural activity Reduced uptake of feed (Spirulina) Reduced movement after disruption Slow valve closure after disruption Increased duration of valve closure Byssal thread attachment inhibition 32 psu 35 psu 45 psu 55 psu 65 psu 75 psu Differential Scale None (-) = little or no change in normal observed behaviour Mild (+) = change in behaviour affecting up to 25% individuals Moderate (++) = change in behaviour affecting up to 50% individuals Severe (+++) = change in behaviour affecting up to 85% or more individuals X.pulex elevated seawater salinity exposure - Pilot study results summary Significant oxygen uptake inhibition occurred at 45 psu and above Significant reduction in digestive gland epithelial layer and functional lumen with increased salinity exposure at 45 psu and above Degradation of interstitial tissue is likely to occur at higher salinity exposures Reduction in valve opening duration with elevated salinity at 45 psu and above Reduction in food uptake with elevated salinity exposure Reduction in byssal thread reattachment at 55 psu and above Higher relative tolerance to increased salinity with gradual salinity increase relative to sudden elevated salinity change Page

148 J. Mondon 1, S. Davis 1 1. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment Deakin University Warrnambool, Victoria Australia 4.3 Seawater hypersalinity stress response - Mytilus edulis sp. Seawater salinity exposure trials were conducted on a larger bivalve species to identify the range in tolerance in a larger species relative to X.pulex, and one which is tolerant of constant submersion. Experimental organism Mytilus edulis sp. Blue mussel Locally available in wild populations and commercially available Wide geographical distribution across southern Australia; from southern WA, across to Victoria and Tasmania and up to northern New South Wales Medium to large in size, and robust to handling; length up to 120 mm Intertidal/shallow sub-tidal species to 15 m depth NB. There has been some confusion as to the name of this particular species. The original species originally thought to be restricted to Australia is Mytilus planulus. The species used could be either M. planulus sp. or M. edulis sp. For the purposes of reporting, M. edulis sp. is used. Figure 9. Mytilus edulis sp. (Linnaeus, 1758) Mytilus edulis sp. (shell length 68.9mm ± 4.3) were obtained from Advance Mussel Supplies (Port Arlington, Victoria), cleaned of epiphytic growth, then transported to Deakin University, Warrnambool, wrapped in damp towels inside an icebox. Five individuals were removed from shell and dissected into <5 mm sections and fixed in 10% formalin to be stored for histological analysis of initial condition. Thirty individuals were placed in 10 x 8 L glass aquaria (tanks) Page

149 containing 4 L of natural seawater and left to depurate for three days. All tanks were housed in an air conditioned facility at ambient air temperature of 12 C, with a natural photoperiod provided by ambient lighting. A 50% water change occurred daily. Water quality parameters were measured every second day and maintained at C ± 0.93, ph 8.05 ±0.09, 32.5 ± 6.89 DO 8.42 mg/l ± Following depuration mussels were fed every two days; the feed rate was 0.02 g dried Spirulina (determined by Spirulina uptake over 12 hours. A rate of 0.02 g Spirulina per individual was maintained throughout testing. Seawater used for husbandry and test procedures was collected from Warrnambool Harbour (38 24 S, E), primary filtered to 20 µm on site, then transported to Deakin University Warrnambool where secondary filtration occurred (4 µm canister filter) prior to storage in sealed aerated containers at 12.4 C ± Salinity levels of stored seawater were increased using Ocean Pure synthetic sea salt (10 g of Ocean Pure added / litre filtered seawater to increase water by 1 salinity unit), and left overnight to allow complete mixing prior to use in experimental treatment tanks. Elevated seawater salinity exposure test summary 1. 3 day laboratory acclimation and depuration 2. Temperature and salinity controlled 3. Ambient daylight photoperiod 4. Filtered natural seawater / artificial sea salts + filtered natural seawater 5. 7 day exposure trial Test treatments Gradual salinity (ramped) exposure was used to investigate the salinity range indicative of maximum brine salinity at discharge. Three 8 litre tanks per salinity exposure were established, with five treatments maintained at the following partial salinity units: 35 (control), 40, 45, 50 and 55 psu. The control salinity value represents in situ salinity from where M. edulis was collected. Mussels were divided into five salinity treatments (5 mussels / treatment / tank), with three replicate tanks per treatment. Gradual salinity increase (ramping) was achieved by 50 % volume water changes conducted daily with elevated salinity seawater to reach the desired final salinity; mussels were then exposed to full treatment salinity for 7 days. Throughout the exposures mussels were fed every second day as described above. Water quality parameters were measured daily and maintained at C ± 1.393, ph 8.05 ±0.09, Page

150 DO 8.42 ± 1.89 mg/l. Nitrite and ammonia levels were monitored at every water change. Test trials were replicated three times (Table 9). Table 9. Experimental design for replicated seawater salinity 7 day exposure trials. Exposure trial SW Salinity (psu) Replicate x2 Replicate x3 Gradual - ramped 35 control 35 control 35 control Statistical analyses Effects of salinity treatments were assessed using a generalized linear model (GLM). Homogeneity of variance of the calculated values was tested using Levene s test, followed by Analysis of Variance (ANOVA). Post-hoc (Tukey s) tests were used to identify where differences occurred between salinity exposures, and against controls. Each biological replicate represents a mean value of three mussels per treatment. All data are represented as mean ± standard error of means (SEM). Differences were considered significant if the probability of Type 1 error was equal to or less than 0.05 (p<0.05). The Effective Concentration of 50% (EC50) for measured endpoints was determined by point estimate (using Logistic (Sigmoidal) curves using log10 transformed rates). The EC50 point estimate value represents the concentration at which test individuals are likely to show a 50% reduction in response rate. Response of M. edulis to elevated salinity exposure was measured across a suite of endpoints. Effect endpoints summary Physiological - metabolic rate (oxygen consumption) Morphological - structural / functional change Behavioural - feeding - movement / byssal thread attachment - gaping Page

151 Metabolic rate - Oxygen Consumption Three individual M.edulis per salinity treatment exposure tank were placed in separate respiration chambers. Chambers were cleaned with 75 C distilled water prior to each trial measurement. Glass lever closure jars (250 ml) were used as respiration chambers, each fitted with a Loligo micro sensor spot, 6 mm glass magnetic stirrer (flea), and protective stainless steel mesh suspended above the base (Figure 10a). A single fibre optic sensor cable per chamber was attached using Velcro straps containing an inbuilt sensor adapter. Chambers were calibrated daily according to Loligo OXY4-Mini transmitter calibration procedure (Loligo 2005). The linear calibration used aerated 100% saturated seawater and 0% saturated solution containing 1g sodium sulphite (AnalaR) in 100 ml seawater. As per the Pilot X.pulex trial methodology, all chambers were submerged in a treatment tank, with three chambers containing one mussel per chamber, and the fourth chamber left unoccupied as a blank. Submerged chambers were left unsealed for 30 minutes to reduce handling stress, then sealed within the treatment tanks, re-attached to the fibre optic sensor cables and moved to the temperature controlled water bath (Figures 10b & c). The water bath was maintained at 12 o C using TempCtrl software (Loligo 2005) which controlled the submersible Ehiem water pump and cooling sump (Figure 4). Oxygen concentration reading output in mg/l was generated by OXY4-mini transmitter software. Each oxygen measurement was conducted over a duration of 1 hour. Shell length, weight and volume were recorded prior to being returned to the respective treatment holding tank. For each test replicate, three individuals (treatment and controls) were sacrificed and fixed in 10% buffered formalin for histological assessment. The methodology described was repeated for all treatments and controls. S. Davis Figure 10. Images showing a) the glass lever closure respiration chamber holding a single mussel, b) the respiration test tank positioned on top of magnetic stirrers, and c) respiration chambers each with the single fibre optic fibre cable attached. Page

152 Results - Oxygen Consumption Statistically significant differences in RR response between elevated seawater salinity exposures is evident (df.=4, F=3.72, p=0.042). Whole animal weight-corrected values represented in Figure 11 and Table 10 indicate the change in oxygen consumption drops significantly at 45 psu, with an upward trend at 55 psu. The effective concentration likely to elicit a 10 % reduction in oxygen uptake is approximately 40 psu, with a 50% reduction in oxygen uptake at approximately 44 psu (Table 11). The increase in consumption at 55 psu is representative of a potential survival spike; when under stress the organism expends energy rapidly to maintain survival over a very short term. Figure 11. Mean ± SE M. edulis weight-corrected oxygen consumption at increasing concentration of seawater salinity exposure. Page

153 Table 10. Comparison between mean weight-corrected oxygen uptake in M. edulis across the exposure salinity range. Bold values indicate the highest oxygen consumption. Letters indicate where statistically significant differences occur across salinity treatments. Salinity Mean respiration Std error Significant rate (mg/l) difference (p<0.05) A A B B AB Table 11. Effective concentration point estimates for oxygen consumption in M. edulis exposed to gradual elevated seawater salinity exposure. Point estimate EC10 EC50 EC90 Oxygen consumption psu psu psu Morphological change - Digestive gland atrophy Mussels were haphazardly selected from each treatment at the end of the treatment exposures. The soft body tissue was removed from the shell by inserting an oyster knife between the valves on the bearded side close to the hinge point at the base of the valves. A scalpel was then used to separate the posterior abductor muscle from the inner shell before removal of the whole animal from the shell. The body was then sectioned into three segments and each immediately fixed in 10% buffered formalin. Fixed samples were dehydrated across a graded series of ethanol from 30% to 100%, cleared in Histolene and impregnated with paraffin wax using a Reichart-Jung Histokinette processor over a 22 hour cycle. Wax impregnated samples were then embedded in paraffin wax using a Tissue-Tek II Embedding Centre. Embedded tissues were sectioned to 5 µm using a HM 325 Microm Microtome, and sections stained with standard Haematoxylin and Eosin (H&E). Page

154 Sections were viewed under a Zeiss Axioplan Universal microscope fitted with a Zeiss HRc digital image capture system. Lauenstein et al. s (1998) semi-quantitative scale was used as a basis to calculate the level of digestive gland atrophy present. Histopathology semi-quantitative scale for digestive gland atrophy. Score Description 0 Normal wall thickness in most tubules (0% atrophy), lumen nearly occluded, few tubules even slightly atrophied 1 Average wall thickness less than normal, but greater than one-half normal thickness, most tubules showing some atrophy, some tubules still normal 2 Wall thickness averaging about one-half as thick as normal 3 Wall thickness less than one-half of normal, most tubule walls significantly atrophied, some walls extremely thin (fully atrophied) 4 Wall extremely thin (100% atrophied), nearly all tubules affected Results Digestive gland atrophy Significant histopathological alteration in the digestive gland of M. edulis was observed in individuals exposed to higher salinity concentrations (df=4, F=4.63, p=0.023) (Figure 12; Table 11). Narrowing of the digestive tubule wall, with corresponding increase in central lumen area, is evident in salinity exposures of psu. Onset of loss of the central star lumen with corresponding decrease in digestive epithelial surface area is clearly demonstrated in digestive tubules from psu exposures (Figure 13). Table 11. Comparison between mean digestive gland (DG) atrophy in M. edulis exposed to change in seawater salinity concentration. Bold values indicate the highest DG atrophy scores. Letters indicate where statistically significant differences occur across salinity treatments. Seawater salinity Mean digestive gland Std error Significant difference (psu) atrophy scale P< C BC B A A Page

155 Figure 12. Mean ± SE M. edulis digestive gland atrophy (scale 0-4) at increasing concentration of seawater salinity exposure. Table 12. Effective concentration point estimates for digestive gland atrophy in M. edulis exposed to gradual elevated seawater salinity exposure. Point estimate EC10 EC50 EC90 Digestive tubule atrophy psu psu psu Page

156 Figure 13. Representative stages of digestive atrophy in M. edulis exposed to increasing seawater salinity concentrations over 7 days. Magnification of images x400. Behavioural change Quantitative and semi-quantitative behavioural observations were made during each exposure trial. Feeding - Feed uptake (semi-quantitative for 3 feeds over 7 days). Feed uptake was calculated semi-quantitatively by imaging the base of the tank and calculating the relative percentage cover of green Spiralina settled on the base of the tank 12 hours post feeding. Feed uptake decreased significantly with increase in seawater salinity exposure at 50 and 55 psu (df. = 4, F=27.85, p<0.0001) (Table 13). Page

157 Table 13. Comparison between feed uptake in M. edulis exposed to change in seawater salinity concentration. Bold values indicate the highest DG atrophy scores. Letters indicate where statistically significant differences occur across salinity treatments. Seawater salinity (psu) Total feed uptake semi-quantitative measure (Maximum uptake scale value = 10) Std error A A A Significant difference B C Byssal thread re-attachment The number of individuals exhibiting byssal thread re-attachment 24 hr post respirometer testing decreased significantly with increase in seawater salinity exposure (df. = 4, F=15.87, p=0.0002) (Table 14). Table 14. Byssal thread re-attachment in M.edulis 24 hours after physical removal and replacement to treatment tank. Bold values indicate the highest DG atrophy scores. Letters indicate where statistically significant differences occur across salinity treatments. Seawater salinity (psu) Mean number of individuals exhibiting byssal thread re-attachment 24 hr post RR measurement Std error A A B BC C Significant difference (p<05) Page

158 Value opening Gaping Daily observations of valve opening per individual were recorded at 9 am and 4 pm daily for the duration of the 7 day exposure trial. The mean total number of individuals exhibiting gaping (ie. valve opening) was significantly different across the treatments; the greatest number of animals exhibiting open valves occurred at 40 & 45 psu (df. = 4, F=6.62, p=0.0072) (Table 15). Table 15. Value gaping in M.edulis over the duration of 7 day exposure to elevated seawater salinity exposure. Bold values indicate the highest DG atrophy scores. Letters indicate where statistically significant differences occur across salinity treatments. Seawater salinity (psu) Mean total number of individuals exhibiting gaping (Max. value = 42) Std error B A A Significant difference C C Results summary - M. edulis elevated seawater salinity exposure Significant oxygen uptake inhibition occurred at 45 psu and above, with evidence of a spike at 55 psu Significant reduction in digestive gland epithelial layer and functional lumen with increased salinity exposure at 45 psu; severity of loss increasing at psu Approximate 37% reduction in food uptake with elevated salinity exposure at 50 psu, increasing to a 63% reduction at 55 psu Onset of significant impairment in oxygen uptake and digestive gland atrophy occurs at approximately 40 psu Inhibition in byssal thread re-attachment at 45 psu, increasing in severity at 55 psu Reduction in gaping activity occurring in animals at 50 psu and above Page

159 J. Mondon 1, S. Davis 1 1. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment Deakin University Warrnambool, Victoria Australia 4.4 Desalination brine stress response - Mytilus edulis sp. Desalination brine exposure trials were conducted on M. edulis to identify the range in tolerance to generated brine salinity. This work was conducted at the National Centre of Excellence in Desalination Australia (NCEDA), Rockingham, Western Australia. The test desalination brine was generated at the NCEDA. Brine exposure trial Methodology summary Farmed Mytilus edulis obtained locally from open water mussel lines Experimentation conducted at NCEDA facility 21 C temperature controlled air and water Shaded 14 hour light photo period Dried Spirulina diet Acclimation over 5 days at 34.5 psu Serial gradient of exposure 7 day exposure trial Mussels (65.47 ± 1.18 mm) were obtained from Blue Water Mussels in Rockingham, Western Australia and transferred to 15 L aerated glass aquaria at the NCEDA laboratories. Mussels were maintained in a 14 hour photo period, at 21 C in temperature regulated seawater collected from the Bunbury region (34.5 psu). Acclimated over 5 days allowed observation and assessment of behaviour indicative of healthy mussels (Figures 14 and 15). A 50% volume water exchange occurred daily. Mussels were fed a diet of dried Spirulina (Table 16); 0.05g / mussel / day. Page

160 S.Davis Figure 14. Photograph showing acclimating mussels in aquaria with cooling system for water, prior to experimental brine trials. S. Davis Figure 15. Photographs taken during acclimation phase showing evidence of normal behaviour in M. edulis; gaping, defecation and byssal thread attachment. Experimental exposure summary 5 experimental brine salinity treatments: 34.5, 37, 40, 45, 54 psu 3 replicate tanks per treatment salinity increase via 25% change / day to final test exposure concentration salinity was adjusted using SWRO brine from the NCEDA pilot plant 7 day test exposure at full treatment salinity 50% water exchange /day salinity, temperature, ph measured /day Nine mussels were placed in each of 15 aquaria which were divided into five salinity treatments 34.5 (control), 37, 40, 45 and 54 psu (three replicate tanks per treatment). Mussels were acclimated for 5 days, to enable byssal thread attached and depuration. Mussels were maintained at 21 degrees with a shaded 14 hour light photo period and fed daily with dried Spirulina (Table 16). Page

161 Note: Three Mussels (24.23 ± 0.34 mm) were added to each tank as a contingency to occasional mortality. Treatment salinities were adjusted using SWRO brine created from the NCEDA pilot plant facilities. Salinity was increased over a series of water changes, each increased the salinity by 25% per day. Mussels were then exposed to full treatment salinity for 7 days (Figure 15). Temperature and ph were recorded daily. Salinity was checked with a refractometer and 50 % water volume was exchanged daily. All tanks were checked for mortalities twice daily; an animal was considered dead when no response to a gentle tap of the shell occurred, in which case the animal was removed immediately. Table 16. Nutritional value of Spirulina (Anthrospira sp.) diet per 100 g dry weight. Page

162 Salinity of respective treatments Salinity (psu) /01/ /01/ /01/ /01/2014 1/02/2014 3/02/2014 5/02/2014 7/02/ psu 45 psu 40 psu 37 psu 34.5 psu (Control) Date Figure 15. Schematic of salinity of each treatment during elevated brine salinity exposure experiment at NCEDA Rockingham, January to February Biological endpoints summary Oxygen consumption used as a measure of metabolic stress response Morphological alteration - Digestive gland atrophy Condition index Oxygen consumption - Testing procedure summary Oxygen consumption measured as decline in O 2 over 60 min in closed respirometry chamber. 2 point calibration performed prior to measurements (0% sol. of 10 g /L sodium sulphite & 100% sol. aerated seawater) Individual mussels placed in open chamber for 30 mins., in treatment tanks, then closed & attached to an OXY 4 mini transmitter. O 2 conc. recorded for 70 mins (1 st 10 mins. discarded) Respiration rates were determined using the following equation: Respiration Rate (mg O₂ / kg / hourˉ ¹) = (Change in Oxygen (mg/l) * Volume of chamber (L)) (Time (hrs.) x Wet tissue weight (g)) Page

163 Oxygen consumption was determined by measuring the decline in oxygen in closed respirometers over 60 mins. Mussels were individually placed in Loligo custom made chambers (Figure 16) for 30 minutes in treatment tanks, then closed and attached to an OXY 4 mini transmitter Figures 17 and 18). Oxygen concentration was recorded for 70 minutes (first 10 minutes discounted) in four chambers, three chambers containing mussels and one blank. A two point calibration was performed prior to measurements on each day, 0 % solution of 10g/L sodium sulphite and 100 % solution of aerated seawater. Figure 16. Open control chamber without mussel (left) and single mussel in open respiration chamber (right) in brine treatment tank prior to oxygen uptake measurement. Figure 17. Images showing closed respiration chamber set up during oxygen uptake measurement. Page

164 Figure 18. Observation throughout oxygen uptake measurement was recorded to ensure mussel gape is maintained. Results - Oxygen consumption Control (34.5 psu) respiration rates were similar to initial measurements prior to salinity exposure increase. Mean respiration rates decreased in elevated salinity treatments but high variability occurred at 37 psu and 54 psu (Figure 19). Increase in respiration rate in 54 psu brine treatment exhibited the bimodal stress response that was evident in M. edulis exposed to 54 psu seawater salinity. It is important to note that >50% mortality occurred at brine exposure of 54 psu which implies that the increase in oxygen uptake recorded does not represent an acclimation or likely survival in the longer term. mg O2/kg/hr Initial Brine Treatment Salinity (psu) Figure 19. Mean respiration rates (± SE) of mussels prior (Initial) to salinity increase and following hypersaline brine exposures (n=3). Page

165 Table 17. Comparison between mean weight-corrected oxygen uptake in M. edulis exposed to elevated desalination brine salinity. Bold values indicate the highest oxygen consumption. Letters indicate where statistically significant differences occur across salinity treatments. Brine salinity (psu) Mean respiration rate Std error Significant difference P<0.05 Initial A A AB B B AB Table 18. Effective concentration point estimates for oxygen consumption in M. edulis exposed to gradual elevated desalination brine salinity exposure. Point estimate EC10 EC50 EC90 ( psu) Respiration rate psu psu psu Morphological alteration Histological assessment summary 4 µm tissue sections stained with haematoxylin and eosin Light microscopy and image capture Digestive gland atrophy scale evaluation Following oxygen uptake measurement, mussels were gently pried open, the posterior abductor muscle severed, and morphometric measures, wet tissue weight (± 0.01 g) and shell length (± 0.1 mm), taken prior to fixation. Tissue fixation Two cross sectional tissue samples (<0.5 mm), encompassing the digestive gland, gut, abductor, gonad and gill section, were fixed in 10% buffered-formalin (appropriate salinity) for histological analysis. Page

166 Tissue processing Fixed tissue (approximately 5 mm x 5 mm) was processed using a Leica ASP300S vacuum tissue processor in an ascending graded series of % ethanol (CSR Ltd), prior to being cleared in Histolene (Grale Scientific Pty Ltd), and mounted in paraffin wax (Leica Microsystems Pty Ltd) at 60ºC prior to sectioning at 4 µm using a HM 325 Microm microtome. Standard Haematoxylin (Amber Scientific Pty Ltd) and Eosin (Amber Scientific Pty Ltd) staining (H&E) was used to stain sectioned tissue prior to mounting with DPX (WWR International Ltd). Histological analysis Histological sections were examined using a Ziess Axioplan Universal microscope fitted with a High resolution Microscopy AxioCam HCr camera. Prevalence of alterations to normal tissue structure was recorded with particular emphasis on the digestive gland. Digestive gland atrophy quantification was determined using semi-quantitative analysis. Morphological analysis of digestive tubules: at least five digestive tubules in five distinct frames at 200 x magnification and recorded using AxioCam digital software. Frames were chosen beginning at the top left of the section and a zigzag pattern was followed to obtain subsequent frames. A semi quantitative assessment of digestive tubule state was assigned using a modified version of the method outlined in (Lauenstien and Cantillo 1998) (Figure 20). Histopathology semi-quantitative scale for digestive gland atrophy. Score Description 0 Normal wall thickness in most tubules (0% atrophy), lumen nearly occluded, few tubules even slightly atrophied 1 Average wall thickness less than normal, but greater than one-half normal thickness, most tubules showing some atrophy, some tubules still normal 2 Wall thickness averaging about one-half as thick as normal 3 Wall thickness less than one-half of normal, most tubule walls significantly atrophied, some walls extremely thin (fully atrophied) 4 Wall extremely thin (100% atrophied), nearly all tubules affected Page

167 Figure 20. Representative stages of digestive atrophy in M. edulis exposed to increasing brine salinity concentrations over 7 days. Image magnification x100. Results Histology Significant histopathological alteration in the digestive gland of M. edulis was observed in individuals exposed to high brine salinity concentrations (df=5, F=4.68, p=0.033) (Figure 21; Table 19). Narrowing of the digestive tubule wall, with corresponding increase in central lumen area, is evident in salinity exposures of psu; the highest decrease in epithelial thickness occurring at 54 psu. The effective concentration likely to elicit a 10 % reduction in digestive gland atrophy is approximately 36.5 psu, with a 50% reduction in epithelial cell thickness at approximately 45 psu (Table 20). Figure 21. Mean ± SE M. edulis digestive gland atrophy (scale 0-4) at increasing concentration of desalination brine salinity exposure. Page

168 Table 19. Comparison between mean digestive gland (DG) atrophy in M. edulis exposed to change in desalination brine salinity concentration. Bold values indicate the highest DG atrophy scores. Letters indicate where statistically significant differences occur across salinity treatments. Brine salinity (psu) Mean digestive gland atrophy scale Std error Significant difference P<0.05 Initial D D C C C B Table 20. Effective concentration point estimates for oxygen consumption in M. edulis exposed to gradual elevated desalination brine salinity exposure. Point estimate EC10 EC50 EC90 Digestive gland psu psu - psu atrophy Mussel condition and mortality Mortality number recorded and individual s data removed from calculation data set Volume of each tested mussel determined by water displacement Morphometric measurement - wet tissue weight and shell length Condition index determined by CI = (wet tissue weight *1000)/(shell length *10) One death occurred in a control tank during the experiment. All other mortalities occurred during the last three days of elevated brine salinity exposure. Weight and condition of animals did not show a statistically significant decline relative to the brine salinity gradient exposure (Table 21), however the values for 45 and 54 psu will be skewed due to the loss of weaker animals through mortality. Page

169 Table 21. Mean mortality, length, weight and condition values in M. edulis exposed to brine over 7 days. Values in bold represent statistically significant differences (p<0.05). Initial 34.5 psu 37 psu 40 psu 45 psu 54 psu Mortality Length (mm) ± SE Weight (g) ± SE Condition ± SE 1 (8.3%) 0 1 (8.3%) 4 (33.3%) 7 (58.3%) 66.0 ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ± 0.67 Brine exposure test summary Oxygen uptake decreased in elevated brine salinity treatments - but high variability (54 psu) 10% reduction in uptake likely to occur at 34.5 psu, 50% reduction likely to occur at 38 psu Digestive gland atrophy increased with elevated brine salinity Digestive gland atrophy 10% likely onset at 36.5 psu, 50% likely onset at 44.5 psu Mortality occurred at higher salinity treatments; 33% at 45 psu, 58% at 54 psu Page

170 4.5 Summary of findings Comparison between M.edulis brine salinity exposure, M. edulis seawater salinity exposure and X. pulex seawater salinity test exposures. Hypersalinity stress was evident in oxygen consumption rates of both species of mussel with elevated salinity exposure treatments. Irrespective of salinity source oxygen consumption declined with increasing hypersalinity ie. tolerance levels were directly related to salinity increase to the point beyond which negative impact occurred. Statistical comparisons cannot be made given the difference between species and temperatures at which each set of experiments were conducted. However, response to elevated salinity in each mussel species follows a generalised trend of decrease in oxygen uptake with increased salinity exposure, pointing to a reduction in metabolic rate. The exception to the trend is the bimodal spike in oxygen uptake by M. edulis at 55 psu, irrespective of salinity being seawater or brine derived. Overall, there is no clear evidence that desalination brine derived salinity stress is greater relative to seawater hypersalinity. High mortalities at 55 psu counter the likely scenario that a peak in oxygen uptake at 55 psu implies longevity, and is most likely to be the opposite. It is not known if the weaker animals exhibited a spike in oxygen uptake prior to death before the 7 day exposure was complete. A similar decrease in digestive capability in mussels exposed to increased salinity levels, irrespective of source, is indicated by atrophy of the epithelial layer in tubules of the digestion gland. M. edulis demonstrates a very close pattern between the effective concentrations value associated with inhibition of oxygen uptake and digestive gland atrophy (Table 22). Page

171 Hypersalinity from desalination brine and artificially modified seawater induced significant change in respiration rate, a surrogate of aerobic metabolism, in both intertidal and sub-tidal mussel species. Oxygen consumption declined with increasing hypersalinity, along with digestive gland deterioration. These changes are not necessarily mutually exclusive. Similarly, mortality was directly related to the level of hypersalinity stress. Marine mussels, as with many other marine invertebrates are osmoconformers, ie. mussels maintain an internal environment that is isosmotic to their external environment. Some regulate the cell volume across a narrow range by utilising free amino acids within the cell (Amende et al. 1980, Bradley 2009), others incorporate isolating body fluids by closing the valves, and periodically testing the external surrounding water prior to fully opening to avoid unfavourable conditions (Bakhmet et al. 2012); a particular advantage enabling bivalves to endure prolonged adverse conditions unlike their univalve (e.g. snails and abalone) counterparts. Cell regulation, whilst effective at lower salinity concentrations would generate a physiological cost to the animal over the longer term. At higher salinities, isolating body fluids by valve closure is again only a short term solution. During valve closure, carbon dioxide (CO2) levels build in the haemolymph (blood), thereby lowering ph and forcing sporadic valve opening (Davenport 1977). At high salinities with extended duration of valve closure, the significant change in ph could elicit morphological and histopathological changes in the organism. The identified histological alteration in mussel digestive gland in both M. edulis and X. pulex signifies a stress response aggravated by elevated hypersalinity. Degeneration of cells and atrophy of the tubular epithelial wall in the digestive gland hinders digestion and absorbance of nutrients (Thompson et al. 1974). Digestive gland atrophy is likely to be influenced by changes to the water content in the gut (Bœuf et al. 2001), causing fluid to pass osmotically from digestive wall cells to the tubule lumen (Randall et al. 2002) resulting in dehydration and potential cell death. Limited osmoregulatory capacity could also account for loss of interstitial tissue noted in the digestive gland of both X. pulex and M. edulis, potentially attributable to osmotic reduction in tissue water content at higher salinities (Shumway 1977). In some cases shedding of basophilic cells into the lumen occurs, causing disintegration the structure of the continuous epithelium cell layer forming the tubule lumen (Garmendia et al. 2011). Destruction of intercellular proteins resulting from salinity change (Koehn et al. 1980), Page

172 also inhibit digestive function through cellular atrophy and ensuing malnutrition, potentially preventing recovery if prolonged or severe. In the context of metabolic efficiency, the decline in oxygen consumption reflects a continuous decline in aerobic scope (aerobic metabolism) as a result of increasing level of stress (i.e. salinity) (Sokolova 2013). Any reduction in respiration rate relative to normal, could be associated with a greater cost of basal metabolism; inhibiting energy reserves for protection and damage repair, reduced intake of food, and inhibition of aerobic pathways producing ATP (Adenosine triphosphate) which is essential in transporting chemical energy within cells for metabolism (Ibid). Respiratory rate reduction is primarily induced by stress, and also a potential survival mechanism to prolong life by temporarily reducing high energy expenditure activity (eg. feeding, movement, reproduction) (Ibid.). Blakeslee et al. (2013) identified increasing hypersalinity resulted in significant decline in oxygen consumption in freshwater bivalves. This study signifies a decrease in physiological function in marine species when exposed to an increasing hypersaline gradient, affirming marine osmoconformers (including other molluscs and invertebrates) are expected to exhibit similar trends of declining oxygen consumption under a hypersaline stress once a critical level or threshold is reached. The results from short term hypersaline exposure trials (derived from desalination brine and artificially elevated seawater - 7 day exposures), collectively provide strong evidence of hypersalinity altering the oxygen consumption capability and ultimately affecting behavioural and morphological health condition of the animals. However, longer term acclimation to hypersaline conditions was not investigated. Consequently it is not clear if, and at what point, recovery activity occurs over the longer term, and whether survival of the most tolerant individuals generates an effectively functioning population over extended hypersaline exposure. Page

173 4.6 References Amende LM & Pierce SK (1980) Free amino acid mediated volumne regulation of isolated Noetia ponderosa red blood cells: control by Ca and ATP. Journal of Comparative Physiology. 138: Bakhmet IN, A. J. Komendantov & A. O. Smurov (2012) Effect of Salinity Change on Cardiac Activity in Hiatella Arctica and Modiolus Modiolus, in the White Sea. Polar Biology 35(1): Blakeslee CJ, Galbraith HS, Robertson LS & St. John White, B. 2013, 'The effects of salinity exposure on multiple life stages of a common freshwater mussel, Elliptio complanata', Environmental Toxicology and Chemistry 32(12): Bœuf G & Payan P (2001). How Should Salinity Influence Fish Growth? Comparative Biochemistry and Physiology C. Toxicology and Pharmacology 130(4): Bradley TJ (2009) Animal osmoregulation, Oxford University Press. Davenport J (1977). A Study of the Effects of Copper Applied Continuously and Discontinuously to Specimens of Mytilus Edulis (L.) Exposed to Steady and Fluctuating Salinity Levels. Journal of the Marine Biological Association of the United Kingdom 57(1): Koehn RK, Bayne BL, Moore MN & Siebenaller JF (1980) Salinity related physiological and genetic differences between populations of Mytilus edulis. Biological Journal of the Linnean Society, 14(3-4): Randall D, Bruggren W & French K (2002) Eckert Animal Physiology: Mechanisms and Adaptations. New York, W.H. Freeman and Co. Shumway SE (1977) Effect of Fluctuating Salinity on Tissue Water-Content of 8 Species of Bivalve Mollusks. Journal of Comparative Physiology 116(3): Sokolova IM (2013) Energy-limited tolerance to stress as a conceptual framework to integrate the effects of multiple stressors. Integrative and comparative biology. 53(4): Garmendia L, Soto M, Vicario U. Yungkul K, Cajaraville MP & Marigomez I (2011) Application of a battery of biomarkers in mussel digestive gland to assess long-term effects of the Prestige oil spill in Galicia and Bay of Biscay: Tissue-level biomarkers and histopathology. Journal of Environmental Monitoring 13: Thompson R, Ratcliffe N. & Bayne B (1974) Effects of starvation on structure and function in the digestive gland of the mussel (Mytilus edulis L.). Journal of the Marine Biological Association of the United Kingdom. 54: Page

174 5. Field evaluation Community structure alteration Investigation of in situ biota associated with desalination brine diffuser discharge Anne Brearley (1) and Julie Mondon (2) 1. Oceans Institute and School of Plant Biology, The University of Western Australia, Crawley, Western Australia 2. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment Deakin University Warrnambool, Victoria Australia Key questions 1. To what extent does brine discharge exposure propagate specific alteration in biological function in benthic community species? 2. Which organism/s appear most vulnerable to desalination brine discharge? 5.1 Introduction This pilot survey of the invertebrate populations present at the Perth Desalination Plant diffuser at Kwinana in Cockburn Sound, Western Australia, was used as a proxy to assess potential effects of brine wastewater discharge at the Southern Seawater Desalination Plant at Binningup, 200 km to the south; the assumption being that biota living on the diffuser riser provides a snapshot of species most likely to be affected by desalination brine discharge. At Kwinana the horizontal outlet component of the diffuser is covered by stone rubble, and the upright discharge riser provides structure and stability to an otherwise mobile sand sheet. Both physical structures facilitate settlement and growth of a variety of marine organisms. Development of this invertebrate community has the potential to retard the disposal of wastewater and destabilize the pipe network, thus to maintain efficiency of the diffuser s outflow, the organic growth around the top of the diffuser risers at both Kwinana and Binningup desalination plants is removed every 12 to 20 months during annual maintenance when the integrity and operation of the diffusers are assessed. From mid-march 2015 to March 2016, access to the Binningup diffuser site was not possible. Alternate biofouling specimens, video footage and static images of the risers and cleaning procedures at Kwinana were made available during regular maintenance conducted by Fremantle Commercial Diving. The survey at Cockburn Sound included examination of fouling biota on 40 outfall risers, but did not include the intake which is considered to be similar in biofouling to the outfall diffuser Page

175 risers (Bertolino pers. com., Fremantle Dive 2016). The 40 riser diffuser array lies in m of water with individual risers placed 4 m apart. Height of risers vary with some sections being 2.8 m high while others are short, only 0.4 m, and situated on or within a rubble mound above the seabed (Figure 1). The riser array extends 500 m offshore with the first riser (#40) at 300 m offshore. (a) (b) (c) (d) Figure 1 Diffuser riser fouling growth at Kwinana February Note the poor visibility (a) Riser #1, 1-2 m high, (b) fouling growth of algae, bryozoans and hydroids with grey-white colonial ascidian Didemnum cf. perlucidum overgrowing a large aggregation of the mussel Mytilus edulis, (c) overgrowth around the discharge opening at the apex of the riser, (d) top of the riser after cleaning, showing the scars of barnacles and a large number of M. edulis including live and dead shells, and algae at the base of the cleaned region. In 2012 the entire outer surface of each fibreglass riser was stripped back to bare surface. In August 2014, a modified cleaning protocol was initiated with divers only removing external growth from the apex and inside aperture of the riser, which was repeated 18 months later in February Thus the samples collected from the lower sections of the diffusers in February Page

176 2016 during cleaning maintenance represent the fouling growth since October 2102, i.e. 40 months. The screening survey aims to document the biota living on and around the diffuser discharge area, and to identify biota tolerant of slightly raised salinity and potentially reduced oxygen concentrations. The investigation also identifies and describes attributes that could be enhancing the marine life in the local area. Water quality Discharge brine concentration should be a minimum of x45 dilution desalination wastewater at the edge of the mixing zone, with a management response initiated if dissolved oxygen of bottom waters (equal to 0.5 m above the seabed) drops to 60% saturation (24 hour running medium) or less in the high and/or moderate protection areas of Cockburn Sound. (Water Corporation Perth Seawater Desalination Plant Ministerial Statement 2015) Salinity The plant is designed to operate continuously, drawing water with an input salinity of 35,000 mg/l to 37,000 mg/l at 16 C to 24 C via the new intake structure. This amounts to under 0.02% of the water in the sound being removed per day, which first passes through a pre-treatment filter to protect the pores of the membranes, before being forced through a spiral wound membrane elements of the RO treatment trains. After treatment, the product is treated with lime, chloride and fluoride before being stored and ultimately being blended with water from other sources and entering the municipal integrated supply system. The filter backwash and concentrate stream is returned to the sound. Although the concentrate flow is about 7% salt, the discharge nozzles are designed to act as diffusers, ensuring that mixed water salinity falls to less than 4% within 50 m of the discharge point. As a result, there will be less than 1% increase in salinity of the receiving waters. (Water Corporation 2016 Water Technology Online 2016 Perth Seawater Desalination Plant, Australia.) Water quality Oxygen In 2013 oxygen saturation (%DOsat) was measured at two depths (0.2 m and 0.5m) above the seabed to the north and south of the discharge area. Small differences in dissolved oxygen concentration (0.9% and 2.7%) and standard deviations (1.22% and 0.95%) at North and South NDSO respectively, continue to suggest that there is minimal difference in dissolved oxygen saturation between the two depths (Water Corporation 2013 NSDO Monitoring in Accordance with PSDP Marine Monitoring and Management Plan 6 th May th June 2013.). The screening investigation followed three aspects; (1) systematic collection of fouling biota and identification of species on representative risers, (2) examination of video footage and Page

177 documentation of the sessile biota on all 40 standing risers to the seabed to establish variability between growth on the riser, and the species collection by the diver, and (3) use of still images extracted from the video to conduct a visual census of invertebrates on the riser before sampling by the diver (enabling comparison with the faunal collection), and after sampling providing a window to hidden and cryptic taxa. 5.2 Methodology The original design was to sample at four diffuser risers (#2, 7, 13, 20) approximately 7 m apart and collect biofouling segments from five quadrats (20 cm x 20 cm) at each riser. After the first sampling at riser #2, the design was reduced to three risers (#2, 13, and 20) and four quadrats to comply with dive time constraints. Sample biofouling material within the quadrat was scraped from the outer surface of the riser and placed in numbered calico bags. Samples representative of biota on the riser were collected at different heights above the seabed, and from different sides of the riser. Sampling operations on three risers (#2, 13, and 20) was filmed by the diver and relayed to the vessel for future examination. Single screen shots from the video footage showed the fouling within the quadrat before and after sampling. Photographs of the quadrats on the seabed allowed comparison between the biota on the riser extending upward through the water column, and on the rubble seabed. Video footage taken by the diver during the final maintenance audit of all 40 diffuser risers and manholes allowed observation of differences between the biota on each riser and on the seabed around the base of the structure, thereby providing a wider view of biota in the area. Samples were pre-examined on the day of collection and frozen. In the laboratory, samples were defrosted, drained, photographed and sorted into major taxa groups. Whole samples were weighed, while damp. Freezing and defrosting tissue resulted in fluid loss and shrinkage from ascidians and bivalves which reduced individual weights, and caused some distortion of softbodied ascidians. Specimens were sorted into species based on morphology, however most of the ascidians were not identified; Ascidiacea and Porifera taxonomy is complex and beyond the scope of this survey. Estimation of abundance varied for different taxa groups. Echinoderms, crustaceans (barnacles, crabs and shrimps), polychaetes (segmented worms), and molluscs (bivalves), solitary ascidians and porifera (sponges) were counted. Colonial ascidians and encrusting Page

178 porifera were noted and wet weighed as an indication of abundance. Individual larger species were weighed, and / or measured. Barnacles and the bivalve Mytilus edulis were recorded separately. Counts of dead M. edulis were compromised as the shells were undergoing decalcification and shattered during collection. 5.3 Results Riser and surrounding seabed community composition Fouling organisms on the diffuser risers and seabed were covered by fine organic and inorganic particles. Examination revealed a range of ascidians growing around and over other individuals probably accounting for some of the dead molluscs and barnacles. Numerous dead shells in various states of fragmentation and erosion were bound into the overgrowth, providing stability and structure to the fouling community and additional habitat. Invertebrates sampling The fouling community was dominated by sessile filter and suspension feeders; filter feeders (ascidians, sponges, polychaetes and some crinoids and holothurians and bivalves) that filter large volumes of water and feed on particles in the water column, and suspension feeders (hydroids, bryozoans, hard and soft corals, jelly fish and barnacles) which feed with tentacles or other structures. Figure 2 Diffuser riser #2 Replicate 1. Compound ascidian Didemnum perlucidum (white overgrowth) covering bivalves Mytilus edulis (solid white arrow left bottom) and barnacles Balanus trigonus (dashed blue arrow) 15 live individuals and 10 dead individuals. Image also shows solitary mussels (yellow arrow), ascidians (dashed white arrow) and non-calcified foliose bryozoan (blue arrow). Page

179 Barnacles Balanus trigonus were the most common species (294 individuals), followed by bivalve molluscs (9 species and 134 individuals), and Ascidiacea (9 species and 91 individuals) (Tables 1-3,4). Bivalve molluscs included Mytilus edulis (blue mussel) (40), Hiatella australis (36), Scaeochlamys livida (scallop) (29), Anomia trigonopsis (jingle shell (14), Malleus meridianus (hammer oyster) (4), and three other small species, Lucinidae sp. (6), Ostrea angasi (1), Musculus (2), and gastropod Serpulorbis siphon (1). Sponges in the samples were small and relatively few in number with only three species present, two encrusting and one spherical Tethys species: diffuser riser #2 with one patch of pink encrusting sponge (5 g) and Tethys (3 g), riser #13 no sponges, and riser #20 with a patch of pink encrusting sponge (29 g) and one small patch of orange encrusting sponge. In contrast, Ascidiacea with 9 species recorded accounted for most of the biomass collected: diffuser riser #2 with 8 taxa and 20 individuals (375 g), diffuser #13 with 4 taxa and 14 individuals (291 g), and riser #20 (with 8 species and 41 individuals (803 g), a total of 75 records. Mussels M. edulis and scallop shells S. livida, like the barnacles B. trigonus, varied greatly in size and age and are likely to be early recruits to clean diffuser riser surfaces representing different spawning events over the last three years. Sheets of the colonial didemnid ascidian covered all three species. Aggregations of M. edulis, which formed a dense mass hanging from the diffuser riser surface, were particularly notable. Hydrozoa (Tubularia), and three species of Bryozoa were also represented in the samples. Watersipora cf. arcuata, with a dark brown-orange or dark red, hard calcareous encrusting form with upright pipe-like structures was most common on riser #20 accounting for 458 g of the sample weight. A second uncalcified bryozoan, maroon in colour, was also present in three of the five quadrats from riser #2 and small colonies of Celleporaria were present on riser #13. Motile species were relatively sparse, predominantly smaller and cryptic, and found amongst the other larger biota. The most abundant motile groups comprised shrimps Alpheidae (6) andcrabs (Xanthidae) - Leptodius sp. (4). The largest motile group were asteroids Coscinasterias muricata (3 juveniles, 30 mm), and associated polychaetes Lepidonotus sp. (3). One goby (chordate Osteichthyes) was found at riser #2. Page

180 Table 1 Taxa present on each diffuser riser, and number species present. (a) Diffuser riser #2. Abundance of records in 5 quadrats. Abundance of Colonial Ascidiacea species (Didemnidae white ), which form large mats over the other fouling biota are however under-represented in this table as are the Cnidaria. The letter P indicates the scattered presence of a species that could not be quantified. P2 indicates two species from the specific phylum or class. Phylum Class/Order Number taxa on three diffuser risers Taxa recorded Diffuser riser #2 Abundance Diffuser riser #2 Porifera Porifera Cnidaria Hydrozoa 1 P Annelida Polychaeta Arthropoda/Crustacea Cirrepedia Decapoda Mollusca Bivalvia Bryozoa Bryozoa 3 1 P Wet weight g Echinodermata Asteroidea Chordata Ascidiacea 9 8 P Chordata Osteichthyes (b) Diffuser riser #13 - Abundance of records in 4 quadrats. Phylum Class/Order Number taxa on three diffuser risers Porifera 3 Cnidaria Hydrozoa 1 Taxa recorded Diffuser riser #13 Abundance Diffuser riser #13 Annelida Polychaeta Arthropoda/Crustacea Cirrepedia Decapoda 2 Mollusca Bivalvia Bryozoa 3 2 P 2 Wet weight g Echinodermata Asteroidea Chordata - Ascidiacea Chordata Osteichthyes 1 Page

181 (c) Diffuser riser #20 - Abundance of records in 4 quadrats. Phylum Class Number taxa on three diffuser risers Taxa recorded Abundance Diffuser riser #20 Porifera Cnidaria Hydrozoa 1 Annelida Polychaeta Arthropoda/Crustacea Cirrepedia Decapoda Mollusca Bivalvia Echinodermata Asteroidea 1 Wet weight gm Chordata - Ascidiacea Bryozoa 3 1 P 3 Table 4 Taxa count of each replicate taken from Kwinana diffuser riser #3, #4 and #20 Page

182 Mytilus Size range The size frequency of M. edulis comprising, 221 individuals in total varied in size. Live shells ranged from 7 85 mm with an average length of 36.6 ±1.29 mm (S.E.), representing regular settlement events post-riser cleaning, illustrating that conditions were generally suitable for survival and growth across years (Figure 3). Mortality of M. edulis was also evident by the numbers of dead shells of this very fecund species; diffuser riser #2 with 26 live and 15 dead, riser #13 with 10 live and 28 dead and riser #20 with 4 live and 12 dead shells. Generally dead juvenile and adult mussels were covered by encrusting ascidians (notably Didemnum cf perlucidum) and appeared to have been or were in the process of being buried and smothered. Similarly, dead barnacles of all sizes and age were also common; riser #2 with 120 live and 25 dead, riser #13 with 131 live and 66 dead, and riser #20 with 42 live and 33 dead. Barnacles B. trigonus, also appeared to have been affected by ascidian overgrowth. Abundance M. edulis Shell Length mm Figure 3 Population structure comprising 221 individuals Mytilus edulis on diffuser risers Kwinana Desalination Plant, February Size range extending from 7 85 mm, with a mean length of 36.6±1.3 (SE.). Additional species associated with the 40 riser diffuser array Video footage greatly enhanced detection of the fauna on risers and on the seabed surrounding the discharge area. Footage showed three major faunal categories: motile taxon in the water column, fish, jellyfish, crabs; motile echinoderms on the seabed, the exception being the crinoids which were also present on the risers; and sedentary ascidians and molluscs on the risers and rubble seabed (Table 3). Page

183 The seabed was covered by rubble of varying size, generally cm. In addition to the rubble and ascidians, the seabed was littered with dead shells, most commonly the mussel M. edulis and scallop S. livida, and a thick layer of turf algae and fine detritus that was easily disturbed effectively reducing visibility. Additionally the turf algae attached to both rubble and ascidians masking the underlying material and features needed for species identification. Diffuser risers ranged in height from m above the seabed, to shorter risers surrounded and buried by a deeper cone shaped mound of rubble. The sand fraction was higher in some areas (# 16, 17, 18 and 31). In addition to images of each individual riser, footage of the manholes, was also informative. Whereas the surrounding rubble bed was dominated by large solitary ascidians, the horizontal surface of the manholes was covered in thick algae. Video footage also revealed a thick overgrowth covering the entire surface area of risers to a depth of approximately 200 mm. Depth or thickness of the fouling community was consistent from bottom to top and around the circumference of the riser,, but slightly wider where it cascaded over the flange and gasket below the riser outlet. Generally defined as algae, the growth was very obvious in the videos although the volume of algae in samples was very low. In contrast, hydroids and un-calcified bryozoa were more obvious in the samples than in the video. The algae appeared to have been fragile thin walled species that could have disintegrated during collection. The less abundant calcareous bryozoan Watersipora, was also present on risers in coral-like patches 10 cm or larger. Pieces were also found on the seabed having fallen off the upright structure and continued to grow on the seabed. A few ascidians, solitary species Phallusia, Polycarpa, Pyura sp. and Herdmania grandis, previously H. momus, and colonial species particularly the white mat-forming D. cf perlucidum, were the dominant groups identified on the video and were observed occupying space on the risers. However these were single observations and much lower than the records from the actual faunal collection: riser #2 with 20 individuals, #13 with 14 individuals and #20 with 41 individuals. Overall, a higher number of ascidians were hidden beneath the algal, hydroid and bryozoa canopy and could not be discerned in the video. Similarly ascidians on the rubble seabed were difficult to identify in the video unless the siphons were open and visible. Some individuals may have been covered with algal turf or disturbed by the diver. Interestingly Porifera sponges were more abundant on video footage than in small colonies in faunal samples. Certainly some of the small pink, orange and violet sponges were present on Page

184 the fouling surface, however the videos also showed larger sponges, all of which appeared to be encrusting, dark in colour, grey black or orange, occupying large patches on the riser and around the rubble base. In support of this observation, images of the risers after collection and cleaning showed similar areas that might be sponges. In contrast, large solitary ascidians and large sponges dominated the seabed community. Clusters of M. edulis were also visible on the risers surrounded by a fouling algae hydroidbryozoa community and often coated with the Didemnid ascidians. Similarly, although M. edulis were visible this belied the number of unseen live shells of various size and age hidden in the canopy. Overall the cluster formations of M. edulis, both dead and live, together with other sessile species clearly supports the observation that the fouling community greatly enhances the total space and attachment sites for other taxa and refuge for motile species. Very large populations of the barnacles Balanus trigonus were also present and similarly difficult to see under the fouling growth. Four octopus were identified in the video in holes near the bottom of risers, however the divers suggest that this number is a conservative estimation of the population, as cavities in the rubble piles were typically constructed by octopus and the aggregations of predominantly M. edulis shell middens at the base of risers is evidence of other individuals in the area. Similarly the motile crustaceans represented by 3 blue swimmer crabs Portunus armatus (previously P. pelagicus), are also common and frequently preyed on by octopus. Given that octopus feed on mussels, the fate of dead mussels evidenced by discarded shells cannot be attributed to poor health, even if this was the case. In contrast to the single species of Asteroidea represented by three records of very juvenile Coscinasterias muricata in faunal samples, an additional five species of echinoderms on the seabed were recorded on video. In all, eleven echinoderms were recorded on video, Asteroidea Anthenea cf australiae (3) and Paranepanthia grandis (4), and 1 each of Holothuroidea Cercodemas anceps, and Echinonoidea Centrostephanus tenuispinaus on the rubble seabed. Crinoidea Comatula purpurea (2) were noted on the thick algal cover near the upper part of the riser, and was also observed on the manhole surface (Table 2). Fishes and other taxa associated with diffuser habitats In contrast to the single species of fish (a blenny) in faunal samples, ten additional fish species were recorded video footage on the seabed around the risers. Butterfish Pentapodus vitta (the Page

185 most abundant species) appeared in actively feeding groups around the risers and were at times attracted by the diver activity, but were absent at risers #16, 17, 20 and 21. Other species appeared to be using the cavities within the fouling community as a retreat;including gobble guts Apogon rueppellii (17), grub fish Parapercis haakei (12), redstriped cardinal fish Apogon victoria (7), wrasse, Labridae sp. (3), western blue box fish Stropiurichthys robustus (2), old wife Ensoplosus armatus (1), Western Australian seahorse Hippocampus angustus (1), fan bellied Leather jacket Monacanthus chinensis (1), and the western red Scorpionicod Scorpaena sumptuosa (1). Other species can also be added to the census of the wider diffuser community and surrounding seabed; large sponges, soft corals (Alcyonacea) recorded by the diver, and fan worms (Polychaeta) Protula sp. and jellyfish viewed on video footage. Using still photographs to identify biota on the diffuser structures Images of the sampling areas were photographed by the diver with great care. However, identification of many organisms was difficult due to turbid water, overgrowth of other species, and camera movement and focus. Larger specimens were easier to identify to Phyla groups e.g. ascidians, but features of others such as the sponges even at the Phyla level were not clear (Figure 3). Photographs of the scraped surface after faunal collection however provided a cross section through the fouling mass, representing a window into the arrangement of biota within the algal overgrowth. The window revealed the extent M. edulis and B. trigonus, sponges, colonial ascidians and other organisms directly attached to the surface of the riser, which was not apparent on the surface view. Populations of B. trigonus on each riser were large and included live and dead shells, and attachment scars, representing numerous periods of settlement and high mortality across all age groups. Many of the dead shells were incased in shrouds of colonial ascidians. Examination of the cleaned surface indicated that B. trigonus had settled directly on the riser surface and were likely to have been some of the first organism to settle. Page

186 Page

187 (a) (b) Figure 3 Diffuser riser #2, 2 m high. Quadrat #1. Before (a) and after (b) cleaning. Northwest side of the riser approx. 0.5 m above the rubble substrate shown in the background. Before sampling (a) (P ) shows rich overgrowth over most of the quadrat (20x20 cm), with a large area of white colonial ascidian Didemnum perlucidum (solid arrow). Photograph (b) (P ) taken after the cleaning shows, the grey background of the diffuser riser, white scars of barnacles (dashed arrow), remnants of the colonial ascidian, and side view of bushy, finely branched algae, bryozoan and hydroid growth against a background of rubble covering the horizontal riser of the diffuser. (a) (b) Figure 4 Diffuser riser #2 Quadrat #2. East side of riser and 0.75 m from seabed, before sampling (a)(p ) and after (b)(p ). Note the Mytilus edulis (arrow) partially covered/ smothered by the white ascidian D. perlucidum, which would have been hidden beneath the bushy overgrowth provisionally Aplidium sp. Page

188 (a) (b) Figure 5. Diffuser riser #2 Quadrat #3. Depth not recorded. Before sampling (a) (P ) and after (b) (P ). Yellow solitary tunicate sea squirt Phallusia (arrow) and small pieces of sponge, clearly evident after cleaning (a) (b) Figure 6 Diffuser riser #2 Quadrat #4 South, 2 m above seabed. (a) Before cleaning (P ) showing predominately algal, bryozoa and hydroid overgrowth. (b) After cleaning (P ) showing barnacle scars, and sea squirt Aplidium to the right of image (arrow), now clearly evident. (a) (b) Figure 7. Diffuser riser #2 Quadrat #5 West, 1 m above seabed (a) (P ) showing white colonial ascidian draped over M. edulis, with non-identified red sponge (arrow), (b) after cleaning (P ) showing scars of barnacles, limited visibility. Page

189 Figure 8 (a) Seabed 1. (P ) Figure 8 (b) Seabed 2. (P ) Figure 8 (c) Seabed 3 (P ) Figure 8 (d) Seabed 4 (P ) Figure 8 (e) Seabed 5 (P ) Page

190 Table 3 Presence and estimated percent cover of species collected from diffuser riser. Table 4 Presence and estimated percent cover of species identified from video post riser cleaning. 5.4 Discussion Biofouling fauna sampling from three separate diffuser risers (13 quadrats) documents the variability of opportunistic fauna on hard substrates at Kwinana. Although some species could not be formally identified, they have previously been observed on other hard surfaces, both natural and industrial, (Brearley pers. comm.). The fouling community reflects three or more years of growth by a suite of species from different phyla each of which contribute to the function of the community (Old pers. comm. 2016). Video examination of all 40 diffuser risers has enabled a broader visualization of biota on the riser surface, identification of the cryptic nature of species within the fouling community, and documentation of the abundance of fauna particularly ascidians and a number of potentially vulnerable large motile taxa (crabs, echinoderms, fish, octopus and nudibranchs) on the rubblecovered seabed beyond the riser structures. Both video and faunal approaches provided evidence that environmental conditions have been suitable for settlement and growth 49 species, 34 of which were collected from three risers, Page

191 with an additional 10 fishes, five echinoderms, plus octopus, crabs, nudibranchs, fan worm, and jellyfish recorded by video living on the risers or on the surrounding seabed. The variety of groups and individual species present indicate that surfaces of diffuser risers can support a diverse community of suspension and filter feeders, grazers, detrital feeders and predators in an otherwise sandy seabed environment. Cockburn Sound is a major shipping port with global connectivity. Historically wooden and steel vessels have used the port since European colonization in Over decades, shipping has transported exotic species as fouling organisms in and on the hulls. At least three species recorded in this survey are introduced; the bryozoa Watersipora arcuta, wide spread across temperate and subtropical regions and possibly a very early introduction from Europe (Gowlett-Holmes 2008), the white colonial ascidian D. cf perlucidum, probably a more recent arrival (Smale et al. 2012) and the scallop Scaeochlamys livida, another recent arrival introduced to southern Western Australia in the 1970s and 80s (McDonald and Wells 2007), and believed to have displaced the native scallop species in Cockburn Sound (Morrison and Wells 2008). The post war period of the 1950s brought development of the port, and continues today, facilitated by dredging of deep channels to improve access for major industries, and construction of the southern causeway to Garden Island that decreased water circulation. Dredging and increased turbidity, restricted circulation and an inflow of industrial pollutants including fertilizers brought a decline in water quality and environmental change, the most notable being the death of seagrass Posidonia sinuosa, primarily due to elevated nutrient levels in the 1970s stimulating algal growth that inhibited seagrass photosynthesis and growth (Cambridge1984 and 1986). While nutrient contamination levels have dropped in recent years, nutrients and consequentially phytoplankton (microalgae) are still abundant. However, filter and suspension feeders can play a role in reducing nutrient-enhanced phytoplankton and epiphytic growth. Large ascidians and sponges associated with diffuser riser biofouling are potentially able to filter the overlying water column daily, however some smaller epiphytic species associated with seagrass (hydroids, bryozoans, polychaetes, amphipods and barnacles), can also have a filtering capacity of the same order of magnitude as larger suspension species, and could be crucial to the health of seagrass communities (Lemmens 1996 a & b). Filtering rates are dependent on water temperature, food availability, flow rates, and the health condition of the animal. Page

192 The larger filter feeding mussels M. edulis and barnacles B. trigonus were abundant and varied greatly in size representing regular settlement events indicative of environmental conditions sustainable for habitation over extended months. However, mussel mortality was high, evidenced by the number of dead shells covered by encrusting ascidians (notably D. cf perlucidum) that potentially smother adult and juvenile mussels. Dense mussel mats appeared to have detached and fallen to the seafloor creating mounds around the base of risers. Whether the detachment was linked to predation from larger predators including the eleven-armed seastar (Coscinasterias muricata), overgrowth by ascidians, poor adhesive substrate or reduction in byssal thread production is not clear. The recorded observations on size variability of M. edulis and B. trigonus, and associated factors affecting settlement, growth and mortality, follow those described by Wilson and Hodgkin (1967) who studied five marine mussels (two temperate species including M. edulis, and three tropical Indo-Pacific species in Cockburn Sound) at Coogee and Kwinana in the 1960s. The three-year experimental study using PVC settlement plates, comparable to the surface of the diffuser riser, reported a very similar sequence of fouling development. Plates were initially colonized by hydrozoans (Tubularia) in June. M. edulis spat settled amongst the hydroids in late August and had covered the PVC surface by September, with a second heavy spat-fall in late October and November. Young mussels grew rapidly throughout the summer and the majority first spawned in July when approximately 11 months old and over 3.0 cm in length. Theoretically, the space provided on settlement plates should only support a small fraction of settled spat to grow to full adult size. However, variability in settlement, growth and mortality year to year was attributed to crowding of young mussels. Mortality was high with a small proportion of the 1960 cohort surviving another season to an age of two 2 years or more, the 2-yr old animals readily distinguished from the younger year class by their barnacle and bryozoan-encrusted shells; a feature observed in the current screening survey. After two years, distinction between the two cohorts was no longer possible. Although longevity was not known precisely Wilson and Hodgkin (1967) considered it unlikely that M. edulis lived more than 2 or 3 years. The largest individual being 9.8 cm on January 21, 1963, aged about 27 months. The processes of reproduction and gonad development were initiated by temperature, beginning at the end of summer when water temperature fell below 21 C and spawning in mid- July occurred at the 14 C winter minimum close to shore (Wilson and Hodgkin 1967). With the current concerns related to increase in sea surface water temperature, it will be interesting Page

193 to follow reproductive activity of M. edulis in the future. Currently the mean surface water temperatures in region have risen to 16 C winter to 24 C summer (Rose et al. 2012). The 1967 study documents populations of M. edulis and B. trigonus present amongst patches of hydrozoa (Tubularia) and overgrowth by ascidians, polyzoans (bryozoans) and alcyonarians (soft corals). Additionally, sponges growing amongst the mussels, partially smothered the mussels (Ibid.). Under such competitive conditions suitable surface substrate for settlement would be limited for mussel larvae. The fouling community associated with the diffuser at Kwinana 2016 was predominantly sessile filter and suspensor feeders comprising ascidians, bryozoa, hydroids, molluscs, and barnacles largely located within the algal/hydroid/bryozoa forest, with only a few motile species present of which three were potential predators, including juvenile asteroids (Coscinasterias muricata - the eleven armed seastar). In contrast the video footage of diffuser risers and surrounding seabed revealed a further five species of echinoderms (7 sea stars, 1 sea cucumber, 2 crinoid, 1 sea urchin), and 11 species of fish recorded feeding and moving through the diffuser area. Blue swimmer crabs (P. armatus), and octopus were also recorded and clearly feeding on the abundant mussels M. edulis. The presence of these additional species indicates that the community is more diverse than that indicated by the faunal samples, and the ecology of the community more complex. The presence of asteroids and echinoids, albeit in very low number, was particularly interesting as echinoderms in general are known to be common within Cockburn Sound and can be sensitive to changes in salinity and water quality (Roberts et al, 2010 quoting Chesher 1971). A final point of interest is the extent of colonization by introduced colonial ascidians dominating very large areas across the riser surface. 5.6 Summary Overview - Use of Kwinana diffuser benthic community as a proxy for Binningup Each of the three riser communities described for Kwinana could be considered a small island on a rubble surface each differing slightly but following similar patterns to the other 37 diffusers at Kwinana. Taken as a whole, they represent small islands in an archipelago surrounded by sandy habitat. The challenge is how to equate the differences and similarities of the Kwinana biota to that of Binningup 200 km to the south. Page

194 Comparison of Sites Kwinana The geography and aspect of the two sites differ markedly. Cockburn Sound, 10,050 ha in area is situated south of the Swan River, the port of Fremantle and metropolitan area of City of Perth. The Sound is 15 km long and 10 km wide. Ten km 2 is protected by Garden Island to the west, extensive sand banks to the north, and Point Peron that restricts water flow from the South, providing a well-protected anchorage and port. The main basin m depth is surrounded by a narrow sill of various width and depth (Southern Flats 2-3 m in the south), with most of the Sound surrounded by narrow sand beaches and large areas of industrial surrounding development including the Perth Seawater Desalination Plant on the now levelled sand dunes, although the fore-dune is intact in places. The natural and modified conditions of the Sound are substantial in comparison to oceanic conditions on Continental Shelf beyond Garden Island. Winds from the south and north are strong and wind driven waves develop in line with the southerly and northern fetch. These waves however in many areas do not disturb the seabed or the fauna, and the larger but less obvious disturbances to the seabed are due to anthropogenic changes; industrial development, barriers, dredging, pollutants and ultimately death of seagrass, and fishing. Industrial structures, wharves, pylons and pipelines, that now occupy much of the shoreline within Cockburn Sound have greatly increased the stable area reef suitable for a range of species in what was historically a sandy and seagrass dominated environment. The artificial reefs covered now by species commonly found on the natural reefs along the coast along with species transported across the globe are thriving in the relatively protected nutrient rich waters. The upright diffusers and surrounding rubble supporting a diverse fouling community of hydroids, bryozoans, barnacles, mussels, ascidians and algae, which adds to the complexity of habitats for motile species. The populations of juvenile and mature individuals indicate that conditions have favoured settlement and growth over several years. Binningup In contrast, geography and oceanography at Binningup is quite different to Kwinana in Cockburn Sound. Facing west with no protective islands or major reef structures, beaches and the seabed receive the full energy of oceanic swells, and massive sand movement erodes the beaches or smothers the low pavement reef and seagrass (Posidonia angustifolia). Page

195 The effects of aspect, extremes of water and sand movement, reef height and rugosity along the Geographe coast were clearly demonstrated by differences in the size, abundance and diversity of biota identified from the 70 km survey in March 2012 which extended from Peppermint Beach 45 km south of Bunbury to Lake Preston South, 13 km north of Binningup. The variety of invertebrates was higher, species more variable and individuals generally larger at sites south of Bunbury where reefs were more extensive, higher and protected, than those to the north within 5 km of Binningup where invertebrates were sparse, many were juveniles and single records of a species. Additionally the more defined reef sites south of Bunbury had a greater cover of large sponges, algae, and corals than northern sites where the pavement reef was very flat and low, subject to burial and affording little protection and attachment sites for sessile and motile organisms. The field surveys in 2012 and 2013 examined the fauna associated with or attached to seagrass and pavement reef. One aspect that has not been investigated is examination of other hard stable structures in the general area. It cannot be assumed that fouling on diffuser risers at Binningup would be as abundant or diverse as those at Kwinana. The swell waves and associated sand movement would almost certainly remove some newly settled larvae, and large individuals could similarly be stripped from the surface in high seas. However some species will effectively settle and grow as evidenced in the 2012 and 2103 field survey report (see Section 2). The coastline at Binningup when compared to Kwinana at Cockburn Sound is constantly subjected to far greater water movement and sand disturbance, that in itself suggests dilution and effective mixing of brine wastewater discharge in situ, ie. mixing and dilution reduces the impact potential of brine release on the resident benthic community. 5.7 References Cambridge, M. L. and A. J. McComb (1984). "The loss of seagrasses in Cockburn Sound, Western Australia. I. The time course and magnitude of seagrass decline in relation to industrial development." Aquatic Botany 20: Cambridge, M. L., A. W. Chiffings, Brittan, C., Moore, L., McComb, A.J.(1986). "The loss of seagrass in Cockburn Sound, Western Australia. II. Possible causes of seagrass decline." Aquatic Botany 24: Chesher R Biological impact of a large-scale desalination plant at Key West, Florida.: In Elsevier Oceanography Series, Vol 2, pp Page

196 Edgar, G.J Australian Marine Life Plants and animals of temperate waters Reed Publishing. ISBN Gowlett-Holmes K A field guide to the marine invertebrates of South Australia Notomares ISBN Lemmens J.W The ecological significance of suspension feeders in seagrass meadows of Western Australia CSIRO Marine Laboratories Lemmens J.W. Clapin, G. Lavery P. and Cary J. 1996(a). Filtering capacity of seagrass meadows and other habitats of Cockburn Sound, Western Australia. Marine Ecology Progress Series 143: Lemmens J. S. Kirkpatrick D. and Thomson P. 1996(b). The clearance rates of four ascidians from Marion Lagoon, Western Australia. Seagrass Biology: Proceedings of an international workshop Rottnest Island, Western Australia th January Eds. Kuo J. Phillips R.C Walker D.I. and Kirkman H. Faculty of Sciences, University of Western Australia, Perth, McDonald J. and Wells F.E Results of a 2007 survey of the Swan River region for four introduced marine species. Western Australian Fisheries and Marine Research Laboratories Fisheries Research Report No Morrison H. and Wells F. E Colonisation of Fremantle Port and Cockburn Sound, Western Australia by the eastern Australian scallop Scaeochlamys livida (Lamarck, 1819). Molluscs Research 28 (2): Oceanica Consulting Pty Ltd Fremantle Ports Kwinana Quay Project Ecosystem Integrity: Ecosystem Component Characterisation. Report No. 560_010/1. Roberts D.A. Johnston E.L. and Knott N.A Impacts of desalination plant discharges on the marine environment: A critical review of published studies. Rose T. H. Smale D.A. and Botting G The 2011 marine heatwave in Cockburn Sound, southwest Australia. Ocean Science Smale 2012 Extreme spatial variability in sessile assemblage development in subtidal habitats off southwest Australia (southeast Indian Ocean) Journal of Experimental Marine Biology and Ecology. 438: Water Corporation 2013 NSDO Monitoring in Accordance with PSDP Marine Monitoring and Management Plan 6 th May th June Water Corporation 2015 Perth Seawater Desalination Plant Ministerial Statement No. 655 & 832 Compliance Assessment Report 1 July 2014 to 30 th June Water Corporation 2016 Water technology Online Perth Seawater Desalination Plant, Australia. Wilson B.R. and Hodgkin E.P A comparative account of the reproductive cycles of five species of marine mussels (Bivalvia: Mytilidae) in the vicinity of Fremantle, Western Australia. Australian and Journal of Marine and Freshwater Research 18: Page

197 6. Field evaluation Bivalve health condition, Kwinana desalination diffuser outfall Screening evaluation of health condition of Mytilus sp. mussels growing on diffuser structures and surrounding seabed at the Perth Seawater Desalination Plant, Kwinana Julie Mondon (1) and Anne Brearley (2) 1. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment Deakin University Warrnambool, Victoria Australia 2. Oceans Institute and School of Plant Biology, The University of Western Australia, Crawley, Western Australia Key question: 1. Is histological alteration a relevant biomarker indicating effect of exposure relevant to alteration in biological function linked to in situ brine exposure? 6.1 Introduction This screening survey of mussels present at the Perth Desalination Plant diffuser at Kwinana in Cockburn Sound, Western Australia, was used as a proxy to assess potential effects of brine on sentinel bivalves in close proximity to wastewater discharge, the assumption being that individuals living on the diffuser riser provide a relative snapshot of likely health condition of individuals most likely to be affected by desalination brine discharge. 6.2 Methodology Mussels were randomly collected from biofouling growth at the Kwinana outfall, February 2016 as described in Section 5. Twenty individuals were removed by hand by diver (Fremantle Commercial Diving), from diffuser ports #2 and #13. These were bagged, brought to the surface and placed in an closed Esky for transportation to shore. Mussel shells were gently prised open, the posterior adductor muscle severed, and the soft body removed from the shell. The body was then immediately sectioned by scalpel into 2 or 3 segments (dependent on body size) and fixed immediately in 10% buffered formalin prior to transportation to Deakin University for tissue processing. Tissues were processed through a graded series of ethanols (30% to 100%), cleared in Histolene, and embedded in 60 C parafin wax. Prepared tissues were then sectioned to 4 µm and stained using standard haematoxylin and eosin method. Tissue sections were viewed using a Zeiss Axioplan Universal microscope fitted with HRc image capture software (see Section 4 for full details on tissue preparation). Histological alterations Page

198 in the digestive gland were scored using quantitative and semi-quantitative scales. A minimum of 30 digestive tubular profiles were recorded per individual from digital images. The lumen area relative to digestive wall profile was calculated (Figure 1), and the overall digestive gland atrophy scale was generated from the mean and range of scores recorded per individual (Table 1). Figure 1 Representative calculation area of digestive tubule wall thickness showing the internal lumen and external digestive tubule regions. Table 1. Histopathology semi-quantitative scale for digestive gland atrophy (based on Lauenstein & Cantillo, 1998). Page

199 6.3 Results Animals from diffuser riser #2 exhibited a digestive gland atrophy range of 1-2, with a mean digestive gland atrophy scale of 1.75 (± 0.16 SE). Animals from riser #13 exhibited a digestive gland atrophy range of 1-4, with a mean scale of 2.5 (± 0.48 SE) (Figure 2) Mytilus sp. digestive tubule atrophy range Atrophy score (0-4) Individual mussel record Figure 2. Spread of digestive tubule atrophy scores for Mytilus sp. collected from diffuser risers at Perth Seawater Desalination Plant, Kwinana. Overall, average tubule wall thickness was less than normal, but greater than onehalf normal thickness. Most tubules showed some degree of atrophy, however some tubules were still normal with a very narrow central lumen surrounded by walls extending into the central lumen, with some almost occluding the lumen passage (see examples represented in Figure 3). In the case of some riser #13 mussels, the wall thickness averaged approximately one-half or 50% of normal wall thickness, whereas others presented evidence of fully atrophied tubule walls (Figure 4). In comparison to the brine exposure values derived from NCEDA desalination brine exposures (see Section 4), the mean atrophy score for riser #2 (1.75) equates to a brine salinity exposure equivalent of 34.5 psu (1.78); the mean score of 2.5 for riser #13 equates to a NCEDA brine exposure equivalent of 37 psu. Overall the stages of atrophy represented across the mussels in situ at this point in time indicate physiological stress has occurred but does not suggest stress beyond recovery for % of live individuals viewed. Page

200 Figure 3. Representative stages of digestive atrophy in Mytilus sp. collected from the field at #2 diffuser port, Kwinana. Image magnification x200. Page

201 Figure 4. Representative stages of digestive atrophy in Mytilus sp. collected from the field at #13 diffuser port, Kwinana. Image magnification x200. Page

202 6.4 Summary and implications of findings A variety of stressors, including hypersalinity, provoke changes to the digestive gland structure beyond the normal range of variability (Lowe et al. 1981, Lowe & Clarke 1989, Vega et al. 1989, Garmendia et al. 2011). Surviving animals at the Kwinana site represent long term exposure and acclimation to water quality conditions present. In comparison to farmed mussels from ambient salinity levels of 34.5 psu (Binningup) constituting a baseline for healthy tissue -level condition, the atrophy scale of Kwinana mussels highlights a state of prolonged but survivable long-term histopathological alteration. Necrotic response in digestive tubule structure is both rapidly induced, as indicated by the findings from the 7 day laboratory brine exposures (see Section 4), and is most likely reversible (Langton 1977, Widdows et al. 1984, Cajaraville et al. 1990, Syasina et al. 1997, Zaldibar et al. 2007). This implies that alternating chemical and environment conditions at Kwinana have not necessarily generated a non-reversable necrotic response in musssels in situ. Kwiniana is not a 'chemically clean' environment in relation to oils and metals given the level of surrounding industrialisation. Elevated metals, eg. copper and cadmium, are known to induce digestive atrophy in bivalves (Vega et al. 1989, Weis et al. 1993, Sheir et al. 2010). Copper release from desalination processing (Roberts & Johnston et al 2010) could potentially be a confounding factor. To elucidate whether the moderate level atrophy identified in mussels is linked to proximity to diffuser brine discharge or a response to the wider environment, or both, requires an assessment of animals across a spatial gradient at increasing distance from the diffuer area. Thus, an induction of response in animals at increasing distance from the diffuser region would indicate an environmental stimulus that extends beyond the water quality derived from diffuser wasterwater release; in this case the alteration in general condition of the digestive gland in mussels would indicate either a moderate environmental stress as opposed to a moderate site specific stress. If no difference occurs across the spatial gradient a combination of stressors is likely. More data at the regional scale are required for tissue-level response in mussels to provide a reliable biomarker of ecosystem health assessment. However the present baseline data generated from the living mussels at Kwinana strongly supports the tenet that these mussels are reactive to representative water quality conditions present at the time of sampling. Without further evidence of mussel health condition in individuals along a spatial gradient of increasing distance from the diffuser structures and which have not been overgrown by other Page

203 sessile organisms, competitive influence from the riser biofouling community as a whole cannot be accounted for and incorporated into the diagnosis of ecosystem water quality and mussel health status. 6.5 References Cajaraville MP, Garmendia L, Orbea A, Werding R, Gomez-Mendikute A, Izagirre U, Soto M,, marigomez I Signs of recovery of mussels' health two years after the Prestige oil spill. marine Environmental research 62 S Garmendia, L., Soto, M., Vicario, U., Yungkul, K., Cajaraville, M.P., marigomez, I Application of a battery of biomarkers in mussel digestive gland to assess long-term effects of the Prestige oil spill in Galicia and Bay of Biscay: Tissue-level biomarkers and histopathology. Journal of Environmental Monitoring 13, Langton RW Digestive rythms in the mussel Mytilus edulis. Marine Biology 41, Lauenstein, G., Cantillo, A., 1998 Sampling and analytical methods of the National Status and Trends Program Mussel Watch Project: Update. NOAA Technical Memorandum NOS ORCA, Vol NOAA/National Ocean Service/Office of Ocean Resources Conservation and Assessment, Silver Spring, MD, Lowe, D.M., Clarke, K.R Contaminant-induced changes in the structure of the digestive epithelium of Mytilis edulis. Aquatic Toxicology 15, Lowe, D.M., Moore M.N., Clarke, K.R Effects of oil on digestive cells in mussels: quantitative alterations in cellular and lysosomal structure. Aquatic Toxicology 1, Roberts, D. A., E. L. Johnston and N. A. Knott (2010). Impacts of Desalination Plant Discharges on the Marine Environment: A Critical Review of Published Studies. Water Research 44(18): Sheir, S. K. and R. D. Handy (2010). Tissue Injury and Cellular Immune Responses to Cadmium Chloride Exposure in the Common Mussel Mytilus Edulis: Modulation by Lipopolysaccharide. Archives of Environmental Contamination and Toxicology 59(4): Syasina, LG, Vaschenko MA, Zhandan PM Morphological alterations in the digestive diverticula of Mizuhopecten yessoensis (Bivalia: Pectenidae) from polluted areas of Peter the Great Bay. marine Environmental Research 44, Vega, M.M., marigomez, I., Angulo, E Quantitative alterations in the structure of digestive cells of Littorina littoria on exposure to cadmium. Marine Biology 103, Weis, P., Weis J.S., Couch, J Histopathology and bioaccumulation in oysters Crassostrea virginica living on wood preserved with chromated copper arsenate. Diseases of Aquatic organisms 17, Widdows, J. (1995). Scope for Growth and Contaminant Levels in North Sea Mussels Mytilus Edulis. Marine Ecology Progress Series 127(1-3): Zaldibar, B., Cancio I, Marigomez I Reversible alterations in epithelial cell turnover in digestive gland of winkles (Littorina littorea) exposed to cadmium and their implcations for biomarker measurements. Aquatic Toxicology 81, Page

204 7. Recommendations Julie Mondon (1), Marion Cambridge (2) and Anne Brearley (2) 1. Centre for Integrative Ecology, School of Life and Environmental Science, Faculty of Science, Engineering and Built Environment Deakin University Warrnambool, Victoria Australia 2. Oceans Institute and School of Plant Biology, The University of Western Australia, Crawley, Western Australia 7.1 Biomarkers of effect. Evidence of biological stress from exposure or changes in environmental conditions can be detected at the structural, biochemical and behavioural levels. The following section outlines recommendations based on the findings from the biomarker and field evaluation investigations on the effect of hypersaline seawater and desalination brine on key marine plant and animal species. A number of the testing procedures documented earlier in this report are technically orientated, requiring specific expertise and equipment. The value of these has been to identify the potential physiological mechanisms causing reduction in the health and functioning of the organism. The other testing procedures documented require training to a technician competency level, and without excessive expenditure. These tests/techniques are feasible for use as regular biomarker screening and monitoring procedures. 7.2 Bivalves - Morphological and tissue-level assessment of health condition Histopathological examination of the digestive gland in mussels provides a sensitive indication for diagnosis of ecosystem health, with particular application for long-term monitoring in situ impacts of wastewater release to the marine environment. Alteration in digestive gland tissuelevel structure affects fitness and should be considered a signal of animal stress. Inhibition of digestive and storage function by reduction in capability indicates physiological impairment. Thinning of the digestive tubule wall constitutes a biomarker representing an inducible response to stressful environmental conditions. Atrophy or epithelial thinning is relatively easily measured using semi-quantitative scoring, and enables detection of environmental stress before the onset of wide-scale mortality. ie. an early warning signal of both impending damage and recovery. - Procedural points 1. To identify the level of health status in animals at the site of interest, an absolute minimum of 20 animals / site are required. An evaluation using the semi-quantitative scaling procedure outlined in section 4 is applicable. Reference to the Effective Page

205 Concentrations Table 2.2 in Section 4.4 of the report provides estimates of salinity at which the severity of response in M. edulis is likely to occur. 2. To identify a change in the direction histopathology alteration requires a baseline against which to compare, ie. is the health condition of the mussels improving, remaining stable, or falling. The baseline can be from the initial / previous histological data collected from the site of interest. 3. To help identify whether the site of interest eg. diffuser, is likely to be responsible for the reduction or level of health condition, requires comparison against 2 or more reference sites (ie. animals collected from sites at considerable distance from any contaminant source), or if these are unavailable / not feasible, translocation of farmed mussels in suspended mussel cages at the site of interest and at points progressively further from the site (ie. a spatial gradient above and below, or radiating out from the source ) is a viable option (Giarratano et al. 2010, Serafim et al. 2011). 4. Caged mussels offer an additional advantage over wild animals growing on structures in that the cages can be pulled up from sites where suspended, potentially without the need for divers, and the mussels can be checked and sampled over regular intervals to identify the time at which alteration occurs. Response to handling and measurement of growth rate is easily recorded, particularly if the translocation animals are farmed and of relatively uniform size. 5. Wild animals growing on diffuser structures offer an advantage in that regular biofouling removal during cleaning of structures provides a readily available supply of mussels. 6. As with any animal, gentle handling is important, particularly during collection and extraction from the shell to avoid tissue damage. Removal of the body from the shell is relatively straight forward and placing into fixative (10% formalin or 70% ethanol) easy. The ratio of tissue to fixative volume must be no greater than 10% to allow for complete fixation to occur. Access to histopathology laboratories is also relatively straight forward universities, vet and human pathology businesses can prepare and stain the slides if needed. 7. Viewing the slides can be undertaken using a standard light microscope. Scoring the atrophy level present the digestive tubules should follow the Digestive Gland Atrophy Scale presented in Table 1 Section Kwinana and reference the representative images presented in Figures 13 & 20, Section 4, and Figures 3 & 4, Section 6). Ideally, taking images of the slides will enable record keeping of the cross tubule profiles for baseline Page

206 comparison and on-going monitoring, although this is not mandatory and careful record keeping over time will generate a valid database of health condition. 7.3 Bivalves - (Physiological - metabolic assessment of health condition) Oxygen consumption measures are an indicator of energy expenditure or metabolic rate, and used as a surrogate measure of aerobic metabolism. Change in metabolic rate signifies the level of energy expenditure by an organism, with alteration in rate indicating either favourable or deleterious environmental conditions. The advantages of this biomarker approach are that: it represents metabolic rate and metabolic stress without needing kill the animal oxygen consumption requires a short test duration of no more than an hour it is easy to test multiple replicates there is low handling stress on the test animals Procedural points 1-5 above are applicable to monitoring metabolic efficiency in field organisms at sites of interest. The measurement procedure for oxygen uptake is more complex but achievable. - Procedural points 1. The Loligo system was trialled throughout this project. This is a commercially available system, designed specifically for respiratory rate evaluation and can be customised to suit the size range of the test organism Setting up the Loligo sytem requires technician level competency (Loligo 2005), but once set up and optimised it is relatively straight forward to use. 3. During the early stages of the project sealed glass vials, Schott bottles and snap-lid preserving jars were used effectively as respiration chambers in place of the more expensive Loligo Respiratory Chambers. 4. Alternative, cheaper methods to measure oxygen uptake can be achieved, using calibrated standard oxygen probes for example. If such methods are used, the initial dissolved oxygen reading and final dissolved oxygen reading will provide a measure of total uptake over the time, which is appropriate for the purposes of the biomarker monitoring if the procedure is standardised. The methodology outlined in Section 4 can be modified and optimised to suit where needed. The volume of the respiration chamber needs to be suitably large to support the respiration of the animal over the duration of one hour, but not too large to challenge the sensitivity of the oxygen Page

207 measure if the oxygen uptake by animals is very low. Sealing and un-sealing respiration chambers to insert probes will require some consideration. 5. Movement of animals from the field to where measurement is conducted will require some preparation. It is preferable that the animals are transported wrapped in wet towel or similar and held in a dark cooler with a lid to reduce light and maintain a relatively stable temperature. Once at the site of testing they can be placed in aerated seawater of similar salinity to the collection site and allowed to stabilise prior to measurement. Mussels that are not gaping or close during the measurement procedure should not be included in the data set. 7.4 Bivalves Considerations for transplantation of mussels for monitoring Transplantation of cultured mussels to the site to monitor water quality conditions is a viable option. In addition to the points made above, the health condition of the transplanted animals is known from the outset any deviation to health condition over time can be linked to alteration in environmental exposure conditions. Farmed animals are used to handling and caging if used. Caged animals and cages themselves can be regularly cleaned to reduce interference from settlement from other species. Caging also reduces likelihood of predation, thereby increasing the value of information related to mortality of animals if it occurs. Cages can be situated along spatial gradients from diffuser outfalls, and reference sites if required. Replication of cages is not problematic. The original project design for Binningup included transplantation of mussels to the Binningup site from the mussel farm at Bunbury. Unfortunately the in situ evaluation was completed due to access to site constraints, but the technique has been effectively applied to multiple organic chemical and heavy metal water quality investigations elsewhere using a variety of commercially available cage designs (Figure 1), and each transplant program testing a suite of mussel biomarkers, including histopathology (Table 1). Page

208 Figure 1 Examples of mussel cage options for deployment in situ. Table 1 Selected examples of caged mussel monitoring applications. Reference (Vidal-Liñán et al., 2014) (Regoli et al., 2004) (Cappello et al., 2013) (de los Ríos et al., 2013) (Marigómez et al., 2013) (Brooks et al., 2012) (Beyer et al., 2013) Mussel Species Exposure time Biomarker analyses Mytilus 4 weeks Chemical analyses galloprovincialis enzyme activity, health condition Mytilus 1 month Trace metals, galloprovincialis cytotoxicity, histological analysis Mytilus 30 days Oil analysis, galloprovincialis oxidative stress, histological analysis Mytilus 21 days Oxydative stress, galloprovincialis histology Mytilus 3 weeks Chemical analyses, galloprovincialis oxidative stress, histological analysis Mytilus edulis 6 weeks Chemical analyses, Immune response, histological analysis Mytilus sp. 3 months Chemical analysis, cytotoxicity, histological analysis Notes on mussel number, cage, deployment time, replication 2 m depth, plastic 1cm mesh housing inside 1m square plastic frame. 100 mussels/site. Replicated at 14 sites. 30 days field exposure. Replicated cages at 2 sites. 200 mussels per cage, 2 replicates, secured by divers at 8 m depth 150 mussels, plastic cylindrical cages, 3 replicates, 1 week deployment. 250 mussels per cage, 1 cage per site, 2 sites. Moored with bottom weight and floating bouy, mussels immersed at 2m depth in a PVC lobster trap. 80 mussels, elasticated nylon mesh bags divided into 10 compartments inside stainless steel cages m depth, 6 sites (50 m, two at 100 m, and 200 m away from discharge point, plus two reference sites. Two stainless steel cages (perforated) deployed at 10 m depth roughly 100 m from Wastewater outfall see Figure 2 below. 30 mussels per cage. Page

209 Figure 2. Design of mussel experiment from (Beyer et al., 2013) 7.5 Bivalves References Beyer J, Aarab, N., Tandberg, A. H., Ingvarsdottir, A., Bamber, S., Borseth, J. F., Camus, L. & Velvin, R Environmental harm assessment of a wastewater discharge from Hammerfest LNG: A study with biomarkers in mussels (Mytilus sp.) and Atlantic cod (Gadus morhua). Marine Pollution Bulletin, 69, Brooks, S., Harman, C., Soto, M., Cancio, I., Glette, T. & Marigomez, I Integrated coastal monitoring of a gas processing plant using native and caged mussels. Science of The Total Environment, 426, Cappello, T., Maisano, M., D'Agata, A., Natalotto, A., Mauceri, A. & Fasulo, S Effects of environmental pollution in caged mussels (Mytilus galloprovincialis). Marine Environmental Research, 91, De Los Rios, A., Perez, L., Oritz -Zarragoitia, M., Serrano, T., Barbero, M. C., Echavarri-Erasun, B., Juanes, J. A., Orbea, A. & Cajaraville, M. P Assessing the effects of treated and untreated urban discharges to estuarine and coastal waters applying selected biomarkers on caged mussels. Marine Pollution Bulletin, 77, Giarratano, E., C. A. Duarte and O. A. Amin (2010). "Biomarkers and Heavy Metal Bioaccumulation in Mussels Transplanted to Coastal Waters of the Beagle Channel." Ecotoxicology and Environmental Safety 73(3): Loligo (2005). "Instruction Manual Oxy-4." Retrieved November, Marigomez, I., Zorita, I., Izagirre, U., Oritz -Zarragoitia, M., Navarro, P., Etxebarria, N., Orbea, A., Soto, M. & Cajaraville, M. P Combined use of native and caged mussels to assess biological effects of pollution through the integrative biomarker approach. Aquatic Toxicology, , Page

210 Regoli, F., Frenzilli, G., Bocchetti, R., Annarumma, F., Scarcelli, V., Fattorini, D. & Nigro, M Time-course variations of oxyradical metabolism, DNA integrity and lysosomal stability in mussels, Mytilus galloprovincialis, during a field translocation experiment. Aquatic Toxicology, 68, Serafim, A., B. Lopes, R. Company, A. Cravo, T. Gomes, V. Sousa and M. J. Bebianno (2011). "A Multi- Biomarker Approach in Cross-Transplanted Mussels Mytilus Galloprovincialis." Ecotoxicology 20(8): Tsangaris, C., Hattzianestis, I., Catsiki, V.-A., Kormas, K. A., Strogyloudi, E., Neofitou, C., Andral, B. & Galgani, F Active biomonitoring in Greek coastal waters: Application of the integrated biomarker response index in relation to contaminant levels in caged mussels. Science of The Total Environment, , Vidal-Linan, L., Bellas, J., Etxebarria, N., Nietto, O. & Beiras, R Glutathione S-transferase, glutathione peroxidase and acetylcholinesterase activities in mussels transplanted to harbour areas. Science of The Total Environment, , Page

211 7.6 Seagrass Biochemical, tissue-level and whole plant response to marine hypersalinity. Seagrass meadows are considered sentinels of coastal degradation (Orth et al., 2006), thus good indicators of impacts to the coastal zone. Change in health condition or abundance indicates potential flow-on effects at a broader ecological scale. Seagrass response to stress associated with changes in salinity begins with physiological adjustments, at the sub-lethal level where the first response involves modification of osmotic potential to avoid desiccation and ion toxicity in order to maintain metabolic processes. Sublethal effects on physiology place extra metabolic demands on the plant that will be reflected in reduced growth. In the longer term, declines in shoot density due to shoot mortality will occur if stress persists. The following points summarise the findings of project investigations in the context of native Western Australian seagrass in situ hypersalinity response: - Salinity range tested under experimental conditions corresponded to the highest salinity (54 psu) for undiluted brine, and 46 psu for 50 % diluted brine recorded at a newlycommissioned desalination plant on the Western Australian open coast. It would however, be unlikely for salinities to remain persistently high with adequate design of a diffuser array discharging brine onto an exposed open coast. Posidonia australis is tolerant of high salinity over several weeks, surviving up to 2 weeks at the highest salinity tested, and more than 6 weeks at the intermediate salinity. This species has the capacity to develop very negative water osmotic and water potentials in the leaves, which counteract the dessicating effects of high salinity in the external medium. - Mortality occurring at the highest salinity did not occur evenly in the test plants, and is likely reflecting differences in allocation to shoots produced on differing orders of branching on the connecting rhizome. In a meadow situation, this would lead to loss of some shoots, and thus reduction in density. Some shoots would survive longer, and, providing the stress was removed in time, would serve as the basis for regrowth and meadow recovery. - There are several indicators and potential biomarkers of sub-lethal salinity stress that could be used in an in situ monitoring program. The most obvious bioindicator of stress conditions is loss of previously robust and healthy seagrass meadows. Baseline knowledge is a pre-requisite to recognizing the loss of meadows. Page

212 Response to in situ stress conditions on surviving seagrass plants such as Posidonia australis can be measured and categorised into three levels of integration. 7.7 Seagrass - Whole organism level, evaluated during field surveys. Severe salt stress would be indicated by blackened patches on the leaves ( salt scorch necrosis ), reduced leaf growth resulting in shorter leaves, and finally lower shoot densities due to death of some of the plants in a meadow. A reduction in root growth / development could be expected. Each marker is relatively straight forward to evaluate and quantify. 7.8 Seagrass - Tissue and cellular level, at which plants counteract the dessicating effects of increased salinity. This includes accumulation of salts, and increasing the concentrations of sugars and amino-acids, so that cells maintain their water content but do not accumulate toxic ionic compound levels. These responses can be detected mostly in detailed experiments involving water relations that require specialised equipment and analyses to determine the salinity and exposure time for the response to develop. Once the responses have been defined by experimental testing for a particular species, it is likely that osmolality, ion concentrations and compatible solute concentrations will provide biomarkers of effect in field collections; sugars, in relation to elevated sucrose, and amino acids, in particular elevated proline and potentially tyrosine). 7.9 Seagrass - Biochemical and biophysical level, salt stress imposes major energy demands on plant tissues and cells, but if osmoregulation can maintain a correct physiological balance (homeostasis), then the energy metabolism (photosynthesis and respiration) of the plant will continue to function. Experiments showed that measures of photosynthesis (chlorophyll fluorescence) did not change for some time after plants were exposed to increased salinity, suggesting that osmoregulation was initially very effective in maintaining the physiological balance of the leaf tissue. There appeared to be a temporal lag in onset of reduction in photosynthesis efficiency in experimental plants. It is unlikely such a lag response would be evident in field organisms. Evaluation of photosynthetic efficiency of seagrass leaves can be reliably determined using a Diving-PAM Underwater Chlorophyll Fluorometer (WALZ Germany) to measure seagrass photosynthesis in situ (Ralph et al. 1998). Page

213 ( The advantage of the submersible PAM over standard field PAMS is that seagrass plants are not disturbed, and repeated measures over extended time scales can be conducted on the same plants / section of meadow Seagrass - Developmental level, seedlings are resilient to brine but exhibit a clear doseresponse to hypersalinity. At 56 psu brine exposure, almost no leaf or root growth was evident for 7 weeks. After recovery in seawater at ambient salinity, shoot growth began within a week. The leaves, however, were pale and stunted, indicating that much of the seed storage reserve (starch and nutrients) had been consumed in sustaining seedlings during their exposure to elevated brine concentration. Leaf growth was present at 47 psu brine concentration but root growth was still markedly inhibited. At 42 psu brine, almost no difference in leaf and root growth compared to seawater controls was evident. Overall, onset of root growth inhibition is likely to occur at concentrations less than required for inhibition of leaf growth. The leaf sheath has been shown to protect expanding young leaves of Posidonia australis (Tyerman 1989) and is likely that this also occurs in germinating seedlings. Seedlings of P. australis are available over a very short period, approx. 2 weeks in early December when the seeds are shed. They exhibit distinct responses to hypersalinity/brine exposure, are easy to culture in controlled conditions and can be assessed in large numbers. Seedling response would therefore provide a useful bioassay to evaluate water quality in relation to hypersalinity during the early life cycle stage. Page